Toxicology of Industrial Compounds
Toxicology of Industrial Compounds Toxicology of Industrial Compounds
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<strong>Toxicology</strong> <strong>of</strong> <strong>Industrial</strong> <strong>Compounds</strong>
<strong>Toxicology</strong> <strong>of</strong> <strong>Industrial</strong><br />
<strong>Compounds</strong><br />
Edited by<br />
HELMUT THOMAS<br />
CIBA-GEIGY Ltd, Basel, Switzerland<br />
ROBERT HESS<br />
Dornach, Switzerland<br />
and<br />
FELIX WAECHTER<br />
CIBA-GEIGY Ltd, Basel, Switzerland
This edition published in the Taylor & Francis e-Library, 2005.<br />
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UK Taylor & Francis Ltd, 4 John Street, London WC1N 2ET<br />
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Copyright © Taylor & Francis Ltd 1995<br />
All rights reserved. No part <strong>of</strong> this publication may be reproduced, stored in a<br />
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the prior permission <strong>of</strong> the copyright owner.<br />
Library <strong>of</strong> Congress Cataloguing Publication data are available<br />
Cover design by Hybert Design & Type, Maidenhead, Berks.<br />
British Library Cataloguing in Publication Data<br />
A catalogue record for this book is available from the British Library.<br />
ISBN 0-203-97962-1 Master e-book ISBN<br />
ISBN 0-7484-0239-X (Print Edition) (cloth)
Contents<br />
Preface vii<br />
List <strong>of</strong> Contributors ix<br />
PART ONE Bioavailability and metabolic aspects <strong>of</strong> industrial<br />
chemicals<br />
1. Biomonitoring and Absorption <strong>of</strong> <strong>Industrial</strong><br />
Chemicals: the Challenge <strong>of</strong> Organic Solvents<br />
F.A.de Wolff S.Kezic J.G.M.van Engelen<br />
A.C.Monster<br />
2. Toxicokinetics and Biodisposition <strong>of</strong> <strong>Industrial</strong><br />
Chemicals<br />
N.P.E.Vermeulen R.T.H.van Welie B.M.de<br />
Rooij J.N.M.Commandeur<br />
3. Metabolic Activation <strong>of</strong> <strong>Industrial</strong> Chemicals<br />
and Implications for Toxicity<br />
G.J.Mulder<br />
4. Sizing Up the Problem <strong>of</strong> Exposure<br />
Extrapolation: New Directions in Allometric<br />
Scaling<br />
D.B.Campbell<br />
PART TWO Reactive industrial chemicals 59<br />
5. Metabolism <strong>of</strong> Reactive Chemicals<br />
P.J.van Bladeren B.van Ommen<br />
60<br />
6. Methods for the Determination <strong>of</strong> Reactive<br />
<strong>Compounds</strong><br />
P.Sagelsdorff<br />
72<br />
PART THREE Pulmonary toxicology <strong>of</strong> industrial chemicals 90<br />
7. Studies to Assess the Carcinogenic Potential <strong>of</strong><br />
Man-Made Vitreous Fibers<br />
T.W.Hesterberg G.R.Chase R.A.Versen<br />
R.Anderson<br />
91<br />
1<br />
2<br />
12<br />
36<br />
44
8. Pulmonary Toxicity Studies with Man-Made<br />
Organic Fibres: Preparation and Comparisons<br />
<strong>of</strong> Size-separated Para-aramid with Chrysotile<br />
Asbestos Fibres<br />
D.B.Warheit M.A.Hartsky C.J.Butterick<br />
S.R.Frame<br />
9. Pulmonary Hyperreactivity to <strong>Industrial</strong><br />
Pollutants<br />
J.Pauluhn<br />
10. Mechanisms <strong>of</strong> Pulmonary Sensitization<br />
I.Kimber<br />
11. Occupational Asthma Induced by Chemical<br />
Agents<br />
C.A.C.Pickering<br />
PART FOUR Biomarkers and risk assessment <strong>of</strong> industrial<br />
chemicals<br />
12. Biomarkers and Risk Assessment<br />
K.Hemminki<br />
13. Extrapolation <strong>of</strong> Toxicity Data and Assessment<br />
<strong>of</strong> Risk<br />
N.Fedtke<br />
14. Molecular Approaches to Assess Cancer Risks<br />
A.S.Wright J.P.Aston N.J.van Sittert<br />
W.P.Watson<br />
15. Evaluation <strong>of</strong> Toxicity to the Immune System<br />
H.-W.Vohr<br />
16. New Strategies: the Use <strong>of</strong> Long-term Cultures<br />
<strong>of</strong> Hepatocytes in Toxicity Testing and<br />
Metabolism Studies <strong>of</strong> Chemical Products<br />
Other than Pharmaceuticals<br />
V.Rogiers M.Akrawi S.Coecke<br />
Y.Vandenberghe E.Shephard I.Phillips<br />
A.Vercruysse<br />
117<br />
129<br />
138<br />
149<br />
157<br />
158<br />
167<br />
180<br />
197<br />
207<br />
PART FIVE Mechanisms <strong>of</strong> toxicity <strong>of</strong> industrial chemicals 222<br />
17. Peroxisome Proliferation<br />
B.G.Lake R.J.Price<br />
223<br />
v
vi<br />
18. Neurotoxicity Testing <strong>of</strong> <strong>Industrial</strong><br />
<strong>Compounds</strong>: in vivo Markers and Mechanisms<br />
<strong>of</strong> Action<br />
K.J.van den Berg J.-B.P.Gramsbergen<br />
E.M.G.Hoogendijk J.H.C.M.Lammers<br />
W.S.Sloot B.M.Kulig<br />
19. Endocrine <strong>Toxicology</strong> <strong>of</strong> the Thyroid for<br />
<strong>Industrial</strong> <strong>Compounds</strong><br />
C.K.Atterwill S.P.Aylward<br />
20. Testing and Evaluation for Reproductive<br />
Toxicity<br />
A.K.Palmer<br />
238<br />
255<br />
280<br />
PART SIX Toxicity <strong>of</strong> selected classes <strong>of</strong> industrial chemicals 300<br />
21. Special Points in the Toxicity Assessment <strong>of</strong><br />
Colorants (Dyes and Pigments)<br />
H.M.Bolt<br />
301<br />
22. <strong>Toxicology</strong> <strong>of</strong> Textile Chemicals<br />
D.Sedlak<br />
309<br />
23. Antioxidants and Light Stabilisers: Toxic<br />
Effects <strong>of</strong> 3,5-Dialkyl-hydroxyphenyl Propionic<br />
Acid Derivatives in the Rat and their Relevance<br />
for Human Safety Evaluation<br />
H.Thomas P.Dollenmeier E.Persohn H.Weideli<br />
F.Waechter<br />
317<br />
24. <strong>Toxicology</strong> <strong>of</strong> Surfactants: Molecular,<br />
Mechanistic and Regulatory Aspects<br />
W.Sterzel<br />
339<br />
PART SEVEN Controversial mechanistic and regulatory issues in<br />
the safety assessment <strong>of</strong> industrial chemicals<br />
25. Low Dose <strong>of</strong> a Genotoxic Carcinogen does not<br />
‘Cause’ Cancer; it Accelerates Spontaneous<br />
Carcinogenesis<br />
W.K.Lutz<br />
26. Controversial Mechanistic and Regulatory<br />
Issues in Safety Assessment <strong>of</strong> <strong>Industrial</strong><br />
Chemicals—an Industry Point <strong>of</strong> View<br />
H.-P.Gelbke<br />
355<br />
356<br />
362<br />
Index 373
Preface<br />
A large number <strong>of</strong> chemical compounds are being constantly introduced<br />
and produced to ease and comfort modern human life. Among those, the<br />
industrial compounds represent that particular fraction <strong>of</strong> chemicals which<br />
are not intended for use in biological systems, but to which humans may be<br />
non-intentionally exposed; at the workplace, by product application or<br />
through the environment.<br />
The International Society for the Study <strong>of</strong> Xenobiotics (ISSX) committed<br />
itself to address, for the first time in the long history <strong>of</strong> industrial<br />
chemicals, the toxicology <strong>of</strong> this class <strong>of</strong> compounds in an intensive<br />
scientific workshop held June 12 through 15, 1994 in Schluchsee,<br />
Germany. This workshop was not only the first such event hosted by ISSX<br />
since its foundation in 1981, but also an extension <strong>of</strong> the society’s scope<br />
beyond its traditionally covered objective to promote studies on xenobiotic<br />
metabolism, disposition and kinetics mainly <strong>of</strong> drugs and agrochemicals.<br />
The large classes <strong>of</strong> pharmaceuticals and agrochemicals had been<br />
deliberately excluded from the scope <strong>of</strong> this workshop, since their terms <strong>of</strong><br />
use generally demand ample registrational toxicity testing that inevitably<br />
leads to a wealth <strong>of</strong> information on, and pr<strong>of</strong>ound toxicological<br />
characterisation <strong>of</strong>, these compounds.<br />
<strong>Industrial</strong> chemicals, instead, which are frequently produced in large<br />
quantities such as pigments, dye-stuffs, plastic materials and additives,<br />
detergents, solvents, etc., to name but a few, are in many cases subjected to<br />
the examination <strong>of</strong> a very basic handling safety only, and may lack any<br />
further toxicity testing. This implies that essentially nothing is known<br />
about their bioavailability, metabolism, excretion and toxicological<br />
properties—unless problems arise. And once toxicity problems come up,<br />
the question arises with them <strong>of</strong> whether or not the available and<br />
traditionally employed methodology is appropriate to approach and solve<br />
them. This, because different from the largely low molecular weight<br />
structures developed for use in biological systems, industrial chemicals are<br />
<strong>of</strong>ten characterised by rather high molecular weight and the incorporation<br />
<strong>of</strong> peculiar structural entities.
viii<br />
Therefore, it was the aim <strong>of</strong> this workshop to contribute to the<br />
investigation <strong>of</strong> industrial chemicals by focussing on the individual<br />
structure, its biological fate, its potential toxicity to mammals and the<br />
molecular mechanisms possibly underlying such adverse effects by<br />
highlighting the use and significance <strong>of</strong> experimental toxicology, with<br />
special emphasis on mechanistic aspects, in the safety assessment <strong>of</strong><br />
industrial compounds as well as to current regulatory and legal<br />
considerations. Topics had been selected to review generally approved facts<br />
and mechanisms, and to particularly address and explore areas <strong>of</strong><br />
investigative and regulatory uncertainty, thereby intending to bring<br />
together the broadly diverse expertise and interests <strong>of</strong> academic<br />
researchers, corporate scientists, experts in safety assessment and<br />
representatives from regulatory authorities.<br />
The following contributions reflect a substantial selection <strong>of</strong> the 27<br />
lectures and six short communications presented during the workshop.<br />
May they succeed in setting a landmark for the due change from the current<br />
era <strong>of</strong> black-box toxicology and largely undifferentiated regulatory<br />
treatment <strong>of</strong> industrial chemicals to the desirable toxicology and safety<br />
assessment by structure in the future.<br />
We gratefully acknowledge the substantial financial support by CIBA-<br />
GEIGY and the RCC Group as well as the financial contributions <strong>of</strong><br />
ADME Bioanalysis, BASF, Henkel, Hüls, Lonza, Schering and Union<br />
Carbide.<br />
Our gratitude is also extended to Mrs Ch.Zehnder for secretarial<br />
assistance and to Taylor & Francis for continuous support, patience and<br />
encouragement to make this publication possible.<br />
H.Thomas<br />
R.Hess<br />
F.Waechter
Contributors<br />
May Akrawi<br />
Department <strong>of</strong> Biochemistry and Molecular Biology, University College<br />
London, Gower Street, London WC1E 6BT, UK<br />
Robert Anderson<br />
Schuller MTC, Health, Safety and Environmental Department,<br />
<strong>Toxicology</strong> Group, PO Box 625005, Littleton, CO 80162–5005, USA<br />
J.Paul Aston<br />
Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK<br />
Christopher K.Atterwill<br />
CellTox Centre, University <strong>of</strong> Hertfordshire, Hatfield Campus, College<br />
Lane, Hatfield AL10 9AB, UK<br />
Samuel P.Aylward<br />
CellTox Centre, University <strong>of</strong> Hertfordshire, Hatfield Campus, College<br />
Lane, Hatfield AL10 9AB, UK<br />
Peter J.van Bladeren<br />
TNO Nutrition and Food Research, PO Box 360, Utrechtseweg 48,<br />
NL-3700 AJ Zeist, The Netherlands<br />
Hermann M.Bolt<br />
Institut für Arbeitsphysiologie, Universität Dortmund, Ardeystrasse 67,<br />
D-44139 Dortmund, Germany<br />
Charles J.Butterick<br />
Texas Technical Health Sciences Centre, Lubbock, TX, USA<br />
D.Bruce Campbell<br />
Servier Research and Development, Fulmer Hall, Windmill Road,<br />
Fulmer, Slough SL3 6HH, UK<br />
Gerald R.Chase<br />
Schuller MTC, Health, Safety and Environmental Department,<br />
<strong>Toxicology</strong> Group, PO Box 625005, Littleton, CO 80162–5005, USA
x<br />
Sandra Coecke<br />
Vrije Universiteit Brussel, Department <strong>of</strong> <strong>Toxicology</strong>, Laarbeeklaan 103<br />
B-1090, Brussels, Belgium<br />
Jan N.M.Commandeur<br />
Division <strong>of</strong> Molecular <strong>Toxicology</strong>, Department <strong>of</strong> Pharmacochemistry,<br />
Vrije Universiteit van Amsterdam, De Boelelaan 1083, H1081 NL-V<br />
Amsterdam, The Netherlands<br />
Peter Dollenmeier<br />
CIBA-GEIGY Ltd., R-1002.2.62, PO Box CH-4002 Basel, Switzerland<br />
Jacqueline G.M.van Engelen<br />
Coronel Laboratory, University <strong>of</strong> Amsterdam, Academic Medical<br />
Centre, Meibergdreef 15, NL-1105 Amsterdam, The Netherlands<br />
Norbert Fedtke<br />
Hüls AG, Bau 2328/PB 12, D-45764 Marl, Germany<br />
Steven R.Frame<br />
DuPont Central Research and Development, Haskell Laboratory, PO<br />
Box 50, Elkton Road, Newark, DE 19714–0050, USA<br />
Heinz-Peter Gelbke<br />
BASF AG, Abt. Toxikologie, D-67056 Ludwigshafen, Germany<br />
Jan-Bert P.Gramsbergen<br />
Department <strong>of</strong> Public Health, Erasmus University, Rotterdam, The<br />
Netherlands<br />
Mark A.Hartsky<br />
DuPont Central Research and Development, Haskell Laboratory, PO<br />
Box 50, Elkton Road, Newark, DE 19714–0050, USA<br />
Kari Hemminki<br />
CNT, Karolinska Institute, Novum, S-141 57 Huddinge, Sweden<br />
Thomas W.Hesterberg<br />
Schuller MTC, Health, Safety and Environmental Department,<br />
<strong>Toxicology</strong> Group, PO Box 625005, Littleton, CO 80162–5005, USA<br />
Elisabeth M.G.Hoogendijk<br />
TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />
NL-2280 HV Rijswijk, The Netherlands<br />
Sanja Keži<br />
Coronel Laboratory, University <strong>of</strong> Amsterdam, Academic Medical<br />
Centre, Meibergdreef 15, NL-1105 Amsterdam, The Netherlands<br />
Ian Kimber<br />
Zeneca Central <strong>Toxicology</strong> Laboratory, Alderley Park, Macclesfield,<br />
Cheshire SK10 4TJ, UK
Beverly M.Kulig<br />
TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />
NL-2280 HV Rijswijk, The Netherlands<br />
Brian G.Lake<br />
BIBRA International, Woodmansterne Road, Carshalton, Surrey, SM5<br />
4DS, UK<br />
Jan H.C.M.Lammers<br />
TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />
NL-2280 HV Rijswijk, The Netherlands<br />
Werner K.Lutz<br />
Universität Würzburg, Institut für Toxikologie, Versbacher Strasse 9,<br />
D-97078 Würzburg, Germany<br />
Aart C.Monster<br />
Coronel Laboratory, University <strong>of</strong> Amsterdam, Academic Medical<br />
Centre, Meibergdreef 15, NL-1105 Amsterdam, The Netherlands<br />
Gerard J.Mulder<br />
Center for Bio-Pharmaceutical Sciences, Sylvius Laboratories, Leiden<br />
University, PO Box 9503, NL-2300 RA Leiden, The Netherlands<br />
Ben van Ommen<br />
TNO Nutrition and Food Research, PO Box 360, Utrechtseweg 48,<br />
NL-3700 AJ Zeist, The Netherlands<br />
Anthony K.Palmer<br />
Huntingdon Research Centre Ltd., PO Box 2, Huntingdon, Cambs,<br />
PE18 6ES UK<br />
Jürgen Pauluhn<br />
BAYER AG, Department <strong>of</strong> <strong>Toxicology</strong>, Institute <strong>of</strong> <strong>Industrial</strong><br />
<strong>Toxicology</strong>, Bldg. 514, D-42096 Wuppertal, Germany<br />
Elke Persohn<br />
CIBA-GEIGY Ltd., Cell Biology Unit, R-1058.2.64, PO Box, CH-4002<br />
Basel, Switzerland<br />
Ian Phillips<br />
Department <strong>of</strong> Biochemistry, Queen Mary and Westfield College,<br />
University <strong>of</strong> London, Mile End Road, London, E1 4NS, UK<br />
C.A.C.Pickering<br />
North West Lung Centre, Wythenshawe Hospital, Southmoor Road,<br />
Manchester M23 9LT, UK<br />
Roger J.Price<br />
BIBRA International, Woodmansterne Road, Carshalton, Surrey SM5<br />
4DS, UK<br />
xi
xii<br />
Vera Rogiers<br />
Vrije Universiteit Brussel, Department <strong>of</strong> <strong>Toxicology</strong>, Laarbeeklaan<br />
103, B-1090 Brussels, Belgium<br />
Ben M.de Rooij<br />
Division <strong>of</strong> Molecular <strong>Toxicology</strong>, Department <strong>of</strong> Pharmacochemistry,<br />
Vrije Universiteit van Amsterdam, De Boelelaan 1083, H1081 NL-V<br />
Amsterdam, The Netherlands<br />
Peter Sagelsdorff<br />
CIBA-GEIGY Ltd., Cell Biology Unit, R-1058.2.52, PO Box, CH-4002<br />
Basel, Switzerland<br />
Dieter Sedlak<br />
Enviro Tex GmbH, Provinostrasse 52, D-86153 Augsburg, Germany<br />
Elizabeth Shephard<br />
Department <strong>of</strong> Biochemistry and Molecular Biology, University College<br />
London, Gower Street, London WC1E 6BT, UK<br />
Nico J.van Sittert<br />
Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK<br />
Willem S.Sloot<br />
TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />
NL-2280 HV Rijswijk, The Netherlands<br />
Walter Sterzel<br />
Henkel KGaA, TTB-Toxikologie, Geb. Z33, D-40191 Düsseldorf,<br />
Germany<br />
Helmut Thomas<br />
CIBA-GEIGY Ltd., Cell Biology Unit, R-1058.2.46, PO Box, CH-4002<br />
Basel, Switzerland. Current address: Ciba-Pharmaceuticals, Stamford<br />
Lodge, Wilmslow, Cheshire SK9 4LY, UK<br />
Kornelis J.van den Berg<br />
TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />
NL-2280 HV Rijswijk, The Netherlands<br />
Yves Vandenberghe<br />
Vrije Universiteit Brussel, Department <strong>of</strong> <strong>Toxicology</strong>, Laarbeeklaan 103<br />
B-1090 Brussels, Belgium<br />
Antoine Vercruysse<br />
Vrije Universiteit Brussel, Department <strong>of</strong> <strong>Toxicology</strong>, Laarbeeklaan 103<br />
B-1090 Brussels, Belgium<br />
Nico P.E.Vermeulen<br />
Division <strong>of</strong> Molecular <strong>Toxicology</strong>, Department <strong>of</strong> Pharmacochemistry,<br />
Vrije Universiteit Amsterdam, De Boelelaan 1083, H1081 NL-V<br />
Amsterdam, The Netherlands
Richard A.Versen<br />
Schuller MTC, Health, Safety and Environmental Department,<br />
<strong>Toxicology</strong> Group, P.O. Box 625005, Littleton, CO 80162–5005, USA<br />
Hans-Werner Vohr<br />
Bayer AG, Fachbereich Toxikologie, Institut für Toxikologie<br />
Landwirtschaft, Friedrich-Ebert-Strasse 217, D-42096 Wuppertal,<br />
Germany<br />
Felix Waechter<br />
CIBA-GEIGY Ltd, Cell Biology Unit, R-1058.2.68, PO Box, CH-4002<br />
Basel, Switzerland<br />
David B.Wahrheit<br />
DuPont Central Research and Development, Haskell Laboratory, PO<br />
Box 50, Elkton Road, Newark, Delaware 19714–0050, USA<br />
William P.Watson<br />
Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK<br />
Hansjörg Weideli<br />
CIBA-GEIGY Ltd, R-1002.2.59, PO Box, CH-4002 Basel, Switzerland<br />
Ronald T.H.van Welie<br />
Division <strong>of</strong> Molecular <strong>Toxicology</strong>, Department <strong>of</strong> Pharmacochemistry,<br />
Vrije Universiteit van Amsterdam, De Boelelaan 1083, H1081 NL-V<br />
Amsterdam, The Netherlands<br />
Frederik A.de Wolff<br />
Coronel Laboratory, University <strong>of</strong> Amsterdam, Academic Medical<br />
Centre, Meibergdreef 15, 1105 Amsterdam, The Netherlands<br />
Alan S.Wright<br />
Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK<br />
xiii
PART ONE<br />
Bioavailability and metabolic aspects <strong>of</strong><br />
industrial chemicals
1<br />
Biomonitoring and Absorption <strong>of</strong> <strong>Industrial</strong><br />
Chemicals: the Challenge <strong>of</strong> Organic Solvents<br />
FREDERIK A.DE WOLFF*, SANJA KEŽI , JACQUELINE<br />
G.M.van ENGELEN and AART C.MONSTER<br />
University <strong>of</strong> Amsterdam, Academic Medical Center,<br />
Amsterdam<br />
Introduction<br />
Organic solvents form a very important group <strong>of</strong> industrial chemicals.<br />
They are widely used in a range <strong>of</strong> occupational settings and may exert a<br />
number <strong>of</strong> deleterious effects when subjects are acutely or chronically<br />
exposed. Among the acute effects are skin and mucosal irritation and<br />
general anaesthesia produced by most solvents at high air concentrations.<br />
Examples <strong>of</strong> chronic effects are peripheral neuropathy after long-term<br />
exposure to n-hexane or carbon disulphide, and the organo-psychosyndrome<br />
or ‘solvent dementia’ which may occur after chronic<br />
occupational exposure to a variety <strong>of</strong> volatile organic compounds.<br />
In order to prevent workers from developing solvent-induced<br />
occupational disease, it is essential to set standards for the duration and the<br />
level <strong>of</strong> external exposure. For a scientifically based standard, a clear<br />
understanding is required <strong>of</strong> the relationship between external exposure,<br />
the uptake by the body, the metabolic fate and the internal dose <strong>of</strong> the<br />
substance. The purpose <strong>of</strong> this contribution is to demonstrate the value <strong>of</strong><br />
biokinetic studies in humans to provide a sound scientific basis for<br />
regulatory decisions on occupational standards.<br />
Biological monitoring<br />
In occupational health practice, monitoring is a tool to protect workers<br />
from developing chemically-induced disease. Monitoring in preventive<br />
health care is described as ‘a systemic continuous or repetitive healthrelated<br />
activity, designed to lead if necessary to corrective action’. In<br />
occupational health, a complete monitoring programme consists <strong>of</strong> four<br />
parts: environmental, biological and biological effect monitoring, and<br />
* Also: University Hospital <strong>of</strong> Leiden, Leiden. The Netherlands
F.A.DE WOLFF ET AL. 3<br />
health surveillance. The latter is a major task for the occupational health<br />
physician, but biological monitoring and biological effect monitoring are<br />
fields <strong>of</strong> interest to the occupational toxicologist. In this contribution, only<br />
biological monitoring will be expounded upon.<br />
Biological monitoring (BM) is defined as the ‘measurement and<br />
assessment <strong>of</strong> workplace agents or their metabolites either in tissues,<br />
secreta, excreta or any combination <strong>of</strong> these to evaluate exposure and<br />
health risk compared to an appropriate reference’ (Zielhuis & Henderson,<br />
1986). This means that a biological monitoring programme is not limited<br />
to the assay <strong>of</strong> xenobiotics in biological samples. As in clinical laboratory<br />
medicine, the pre-analytical phase <strong>of</strong> the process is very important, and<br />
even more so the post-analytical phase <strong>of</strong> the laboratory analysis, which<br />
means the interpretation <strong>of</strong> the analytical data in biomedical terms. The<br />
ultimate goal <strong>of</strong> biological monitoring is the evaluation <strong>of</strong> the health risk <strong>of</strong><br />
workers by estimation <strong>of</strong> the internal dose <strong>of</strong> a chemical. This is not limited<br />
to measurement <strong>of</strong> the quantity <strong>of</strong> the substance absorbed by the body, but<br />
may also include the assay <strong>of</strong> metabolites <strong>of</strong> toxicological interest, if<br />
possible in or near a critical organ (Monster & van Hemmen, 1988).<br />
This implies that the absorption, metabolism and elimination <strong>of</strong> a<br />
substance in man should be known before a biological monitoring<br />
programme can be performed in practice. Animal experiments are <strong>of</strong><br />
limited value; volunteer studies in order to determine pulmonary and<br />
dermal uptake <strong>of</strong> organic solvents provide more relevant data for this<br />
purpose.<br />
Owing to the existence <strong>of</strong> very sensitive analytical methods it is possible<br />
to study the kinetics and metabolism <strong>of</strong> solvents in volunteers who are<br />
experimentally exposed to levels at or far below the <strong>of</strong>ficial threshold limit<br />
values, so that any health risk for the volunteers can almost totally be<br />
excluded.<br />
As with biological monitoring <strong>of</strong> most other substances, in the case <strong>of</strong><br />
organic solvents the compound itself and/or its metabolite in blood or urine<br />
can be measured. Studies with volatile, rather lipophilic, substances have an<br />
additional advantage, namely that the solvent can also be measured in<br />
expired air. Analytically this has the advantage <strong>of</strong> an extremely clean<br />
matrix in comparison with body fluids, whereas biologically, air samples<br />
provide us with information on the blood concentration <strong>of</strong> a volatile<br />
compound. Moreover, collection <strong>of</strong> expired air is non-invasive and large<br />
volumes are readily available (Droz & Guillemin, 1986).<br />
An example <strong>of</strong> a study on solvents in volunteers is the one carried out in<br />
our laboratory on the biokinetics <strong>of</strong> n-hexane and its neurotoxic metabolite<br />
2,5-hexanedione (Van Engelen et al., in preparation). Volunteers are<br />
exposed during 15 min to 60 ppm hexane by inhalation. The minute volume<br />
and the respiratory rate are measured and blood and exhaled air sampled<br />
frequently for determination <strong>of</strong> 2,5-hexanedione and n-hexane,
4 BIOMONITORING AND ABSORPTION OF INDUSTRIAL CHEMICALS<br />
respectively. Each volunteer is exposed twice in succession on one test day<br />
in order to get an impression <strong>of</strong> the within-day intra-individual variation.<br />
Venous blood is sampled through a catheter, and alveolar air is collected<br />
after holding breath for 30 s (to achieve equilibrium between pulmonary<br />
blood and air) by exhaling through a glass tube which is stoppered<br />
immediately. These tubes contain 70 ml alveolar air and the total volume is<br />
analyzed for n-hexane by using a purge-and-trap system. 2,5-Hexanedione<br />
in serum is measured by using electron capture detection after<br />
derivatization, with a detection limit <strong>of</strong> 30 micro-mol l −1 (Keži and<br />
Monster, 1991). During exposure the concentration <strong>of</strong> n-hexane in<br />
alveolar air increases very rapidly and decreases after discontinuation <strong>of</strong><br />
exposure. The half-life time <strong>of</strong> exhalatory elimination after the distribution<br />
phase is in the order <strong>of</strong> 30 min.<br />
2,5-Hexanedione becomes detectable in blood as fast as 2–3 min after<br />
commencement <strong>of</strong> n-hexane exposure. After discontinuation <strong>of</strong> dosing the<br />
metabolite concentration continues to increase for another 3 min, to<br />
disappear from the plasma with a half-life <strong>of</strong> approximately 1.5 h. The<br />
second exposure period on the same day shows very reproducible n-hexane<br />
and 2,5-hexanedione curves in the same individual. Between individuals<br />
there is considerable variation in kinetics and metabolism, and this issue is<br />
being studied in detail at present.<br />
Before a biological monitoring programme can be designed, a detailed<br />
biokinetic study like this one, <strong>of</strong> every solvent being used in industry, has to<br />
be performed. Without kinetic data it is impossible to choose for instance<br />
the correct matrix, the compound to be measured, or the sampling<br />
frequency and time. In addition, these data are necessary to establish a<br />
relationship between ambient air concentrations <strong>of</strong> a chemical (external<br />
exposure), and the biological parameters used to estimate a health risk.<br />
Absorption<br />
The primary association <strong>of</strong> the pharmacologist or general toxicologist,<br />
when reading or hearing the term ‘absorption’, is with ‘intestinal’. For<br />
drugs, gastrointestinal uptake is indeed the most common route to enter<br />
the body. In case <strong>of</strong> occupational exposure, however, intestinal absorption<br />
is <strong>of</strong> minor importance. The occupational toxicologist is, therefore, more<br />
inclined to pay attention to entry routes other than the intestine, the most<br />
important being pulmonary and dermal uptake.<br />
Pulmonary uptake<br />
There are a number <strong>of</strong> parameters which affect the pulmonary uptake <strong>of</strong><br />
organic solvents. In the first place, the physical chemistry <strong>of</strong> the compound<br />
is <strong>of</strong> importance. Both the blood-to-gas and the tissue-to-blood partition
F.A.DE WOLFF ET AL. 5<br />
Figure 1.1 The mean minute volume (1 min −1 ) and the percentage <strong>of</strong> the minute<br />
volume cleared from solvent (shaded area) during exposure to styrene (left) and 1,1,<br />
1-trichloroethane (right) at increasing degree <strong>of</strong> workload.<br />
coefficients determine the absorption through the alveolar membrane and<br />
the distribution over the body. Furthermore, exercise is an important<br />
physiological determinant. With increasing exercise, ventilation increases<br />
and, therefore, also the availability <strong>of</strong> the vapour to the lung per unit <strong>of</strong><br />
time. In addition, cardiac output increases during exercise, and this may<br />
affect absorption, distribution and metabolism through enhanced blood<br />
flow.<br />
Finally, the elimination <strong>of</strong> a solvent which occurs during exposure may<br />
significantly affect the uptake rate. The percentage <strong>of</strong> the vapour not<br />
retained by the body but exhaled again is dependent on, again,<br />
physicochemical factors such as solubility, but also on the rate <strong>of</strong><br />
metabolism (Fiserova-Bergerova, 1985).<br />
In order to demonstrate the different factors which may affect<br />
pulmonary absorption <strong>of</strong> vapours we have constructed Figure 1.1, based<br />
on earlier work <strong>of</strong> Astrand et al. (Astrand, 1975). In their studies,<br />
volunteers were exposed to different vapours such as styrene or 1,1,1trichloroethane<br />
at increasing degrees <strong>of</strong> workload during 2 h.<br />
The first 30 min they were exposed at rest, and then the workload was<br />
increased every 30 min with 50 W. The minute volume, here referred to as<br />
‘supply’, was measured and expressed in 1 min −1 , and the exhaled solvent<br />
concentration was also measured at regular intervals. The shaded area <strong>of</strong><br />
the vertical bars in Figure 1.1 indicate the percentage <strong>of</strong> minute volume<br />
cleared from the solvent, averaged over the observation period. This is<br />
considered to be a measure for pulmonary uptake.
6 BIOMONITORING AND ABSORPTION OF INDUSTRIAL CHEMICALS<br />
During continuous exposure to a constant concentration and at<br />
increasing exercise the uptake <strong>of</strong> styrene remains constant, expressed in<br />
terms <strong>of</strong> percentage <strong>of</strong> the minute volume cleared. Apparently, the body is<br />
not easily saturated with styrene. The picture for 1,1,1-trichloroethane is<br />
completely different. Although the minute volume at each level <strong>of</strong><br />
workload is comparable with that <strong>of</strong> the styrene experiment, it is clear that<br />
the retention <strong>of</strong> 1,1,1-trichloro-ethane is much lower. Apparently, the body<br />
becomes rapidly saturated with 1,1,1-trichloroethane. The reasons for the<br />
difference in pulmonary uptake between these two solvents are evident.<br />
Styrene is highly soluble in blood and it is extensively metabolized to<br />
mandelic acid and phenyl glyoxylic acid. The retention in the body remains<br />
the same, and therefore the uptake increases proportionally with the<br />
minute volume.<br />
In contrast to styrene, 1,1,1-trichloroethane has only a limited solubility<br />
in blood, and it is hardly metabolized. This means that during exposure the<br />
body becomes rapidly saturated with the substance, and that an increase in<br />
minute volume by increasing workload results in a lower retention, and<br />
hardly in higher uptake. Differences in kinetic behaviour, as demonstrated<br />
for styrene and 1,1,1-trichloroethane, are important for the design <strong>of</strong> a<br />
biological monitoring programme.<br />
Dermal uptake<br />
Absorption <strong>of</strong> solvents through the skin may be affected by a number <strong>of</strong><br />
factors. Many organic solvents are able to penetrate the skin and thus enter<br />
the body. This is a rather well-known fact which can be prevented in<br />
industrial practice by use <strong>of</strong> protective clothing. It is, however, less<br />
common knowledge that solvents in the vapour phase may also penetrate<br />
the skin. In case <strong>of</strong> skin exposure to liquids usually a small surface is<br />
exposed, whereas in case <strong>of</strong> vapour the whole body surface <strong>of</strong> about 2 m 2<br />
may be exposed. This means that under certain conditions skin absorption<br />
<strong>of</strong> vapour may significantly contribute to the amount absorbed by<br />
inhalation.<br />
Other parameters which may affect skin absorption are the temperature,<br />
and the ability <strong>of</strong> some solvents to increase their own absorption by<br />
causing skin hyperaemia through irritation. To demonstrate these factors,<br />
some preliminary results are shown <strong>of</strong> a volunteer study on skin<br />
penetration <strong>of</strong> solvents in the liquid and vapour phases (Keži et al., in<br />
preparation).<br />
The experimental conditions are as follows. The volunteer is seated in a<br />
clear-air cabin in order to avoid additional inhalatory exposure to vapour<br />
in the experimental room. The arm is the only part <strong>of</strong> the body outside the<br />
cabin. In case <strong>of</strong> exposure to liquid on the skin, the solvent is put in a<br />
chamber which is pressed on to the skin during the exposure period, which
F.A.DE WOLFF ET AL. 7<br />
is usually no longer than a few minutes. The exposed area is usually in the<br />
order <strong>of</strong> 20 cm 2 .<br />
In the case <strong>of</strong> dermal exposure to vapour, the volunteer places the lower<br />
arm into a piece <strong>of</strong> drainage pipe through which the vapour is led with<br />
controlled flow and concentration in air. Uptake <strong>of</strong> liquid or vapour is<br />
measured in both cases by determination <strong>of</strong> the solvent in expired air, by<br />
the sampling method described earlier.<br />
Figure 1.2 shows the dermal uptake and elimination <strong>of</strong> two different<br />
liquids in one volunteer. A surface <strong>of</strong> 27 cm 2 was exposed during 3 min to<br />
pure 1,1.1-trichloroethane and to tetrachloroethene. It is clear that 1,1,1trichloroethane<br />
is absorbed through the skin much faster and to a much<br />
greater extent than tetrachloroethene, at least in exposure to the liquids.<br />
However, when the skin is exposed to the same solvents in the vapour<br />
phase the picture becomes totally different. Here the lower arm, which has<br />
a surface <strong>of</strong> about 500 cm 2 , was exposed during 15 min to solvent<br />
concentrations <strong>of</strong> approximately 500 µmol 1 −1 air (Figure 1.3).<br />
In the case <strong>of</strong> vapour exposure no difference in absorption kinetics is<br />
observed, and only a small difference in expired air concentration is seen.<br />
The reason for the discrepancy between vapour exposure is that 1,1,1trichloro-ethane<br />
causes skin irritation as the liquid, but not in the vapour<br />
phase. Irritation leads to hyperaemia and, hence, increased absorption.<br />
As it is known that dermal exposure to vapour may lead to detectable<br />
absorption, the contribution <strong>of</strong> vapour uptake <strong>of</strong> the skin in comparison to<br />
inhalatory absorption should be evaluated. This was done with<br />
trichloroethene as an example (Figure 1.4). Both curves were obtained in<br />
the same volunteer. The dermal exposure was performed first, followed by<br />
the inhalatory test after a wash out period <strong>of</strong> 2 weeks. The exposure period<br />
was 15 min, and the inhalatory concentration was 4.1 µmol l −1 . Dermal<br />
exposure <strong>of</strong> the lower arm took place at 1.4 mmol l −1 .<br />
It appears that uptake from the lungs occurs much faster than via the<br />
skin. This is conceivable because the stratum corneum is a stronger barrier<br />
than the alveolar epithelium, and causes a shift to the right <strong>of</strong> the t max. It<br />
can also be seen that inhalatory exposure leads to a much higher expired<br />
air concentration than dermal exposure. But in this respect we should<br />
realize that only a small part <strong>of</strong> the skin was exposed, namely about 500<br />
cm 2 . In fact the result should be extrapolated to the total surface <strong>of</strong> the<br />
human skin, which is about 2 m 2 . These results indicate that dermal<br />
exposure to solvent vapour should not be neglected when the safety <strong>of</strong> the<br />
industrial environment is evaluated. This is <strong>of</strong> special importance when<br />
ambient air concentrations are high, and workers are protected with<br />
protective masks but not with gloves. Another example in which skin<br />
absorption may be high in comparison with inhalation are those solvents<br />
which are readily absorbed by the skin, such as 2-butoxyethanol (Johanson<br />
and Boman, 1991).
8 BIOMONITORING AND ABSORPTION OF INDUSTRIAL CHEMICALS<br />
Figure 1.2 Elimination <strong>of</strong> 1,1,1-trichloroethane and tetrachloroethene by expired<br />
air after dermal exposure to the liquid <strong>of</strong> 27 cm 2 fore-arm skin during 3 min. 1,1,1trichloroethane<br />
liquid irritates the skin.<br />
Figure 1.3 Elimination <strong>of</strong> 1,1,1-trichloroethane and tetrachloroethene by expired<br />
air after dermal exposure to the vapour <strong>of</strong> 500 cm 2 lower-arm skin during 15 min<br />
to 500 µmol l −1 air.<br />
The temperature <strong>of</strong> the solvent is another factor that may have an<br />
influence on uptake through the skin. Figure 1.5 shows the results <strong>of</strong><br />
dermal exposure to liquid tetrachloroethene and n-hexane at two different<br />
temperatures in one volunteer. Exposure time here was only 1 min, and<br />
absorption and elimination were measured by analysis <strong>of</strong> the vapours in<br />
expired air.
At the low temperature <strong>of</strong> the liquid (15°C), the uptake <strong>of</strong><br />
tetrachloroethene is negligible when compared with a normal skin<br />
temperature <strong>of</strong> 33°C. In case <strong>of</strong> n-hexane, under comparable circumstances<br />
and in the same volunteer, the effect <strong>of</strong> temperature is much less<br />
pronounced. Apparently, the physicochemical properties <strong>of</strong> the solvent are<br />
an additional determining factor. The mechanism on which the difference<br />
between tetrachloroethene and n-hexane is based is the subject <strong>of</strong> further<br />
study.<br />
Conclusions<br />
F.A.DE WOLFF ET AL. 9<br />
Figure 1.4 Elimination <strong>of</strong> trichloroethene by expired air during and after inhalatory<br />
exposure to 4.1 µmol l −1 trichloroethene during 15 min, and after dermal vapour<br />
exposure during 15 min <strong>of</strong> the lower-arm skin (500 cm 2 to 1.4 mmol l −l air).<br />
In occupational health practice, the major absorption routes for organic<br />
solvents are not ingestion, but inhalation and skin penetration, the latter<br />
both as liquid and as vapour. The physical chemistry <strong>of</strong> the compound,<br />
exercise, and the elimination rate may affect pulmonary uptake. Factors<br />
affecting dermal uptake are the ability <strong>of</strong> the solvent to penetrate the skin as<br />
liquid or vapour, the temperature <strong>of</strong> the liquid, and the irritability <strong>of</strong> the<br />
chemical to the skin.<br />
Before a biological monitoring programme for solvent exposure can be<br />
set up, the kinetics and metabolism <strong>of</strong> the various solvents in man should<br />
be known. Owing to the availability <strong>of</strong> sensitive analytical methods it is<br />
usually possible to perform volunteer studies at safe exposure levels.<br />
Measurement <strong>of</strong> solvents in expired air and <strong>of</strong> their metabolites in body<br />
fluids is <strong>of</strong> the utmost importance to estimate the internal dose <strong>of</strong> the<br />
solvents and health risk to which man can be exposed in the work and<br />
general environment.
10 BIOMONITORING AND ABSORPTION OF INDUSTRIAL CHEMICALS<br />
Figure 1.5 Elimination <strong>of</strong> tetrachloroethene and n-hexane by expired air after<br />
dermal exposure during 1 min to liquid at 15°C and 33°C<br />
References<br />
ǺSTRAND, I., 1975, Uptake <strong>of</strong> solvents in the blood and tissues in man. A review,<br />
Scand J Work Environ Health, 1, 199–218.<br />
DROZ, P.O. and GUILLEMIN, M.P., 1986, Occupational exposure monitoring<br />
using breath analysis, J Occup Med, 28, 593–602.
F.A.DE WOLFF ET AL. 11<br />
FISEROVA-BERGEROVA, V., 1985, Toxicokinetics <strong>of</strong> organic solvents, Scand J<br />
Work Environ Health, 11, suppl. 1, 7–21.<br />
JOHANSON, G. and BOMAN, A., 1991, Percutaneous absorption <strong>of</strong> 2butoxyethanol<br />
vapour in human subjects, Br J Ind Med, 48, 788–92.<br />
KEŽI , S. and MONSTER, A.C., 1991, Determination <strong>of</strong> 2,5-hexanedione in urine<br />
and serum by gaschromatography after derivatization with O-<br />
(pentafluorobenzyl)-hydroxylamine and solid-phase extraction, J Chromatogr,<br />
563, 199–204.<br />
MONSTER, A.C. and VAN HEMMEN, J.J., 1988, Screening models in<br />
occupational health practice <strong>of</strong> assessment <strong>of</strong> individual exposure and health<br />
risk by means <strong>of</strong> biological monitoring in exposure to solvents, In Notten,<br />
W.R.F., Herber, R.F. M., Hunter, W.J. et al. (Eds) Health Surveillance <strong>of</strong><br />
lndividual Workers Exposed to Chemical Agents, pp. 47–53, Berlin: Springer.<br />
ZIELHUIS, R.L. and HENDERSON, P.Th., 1986, Definitions <strong>of</strong> monitoring<br />
activities and their relevance for the practice <strong>of</strong> occupational health, Int Arch<br />
Occup Environ Health, 57, 249–57.
2<br />
Toxicokinetics and Biodisposition <strong>of</strong> <strong>Industrial</strong><br />
Chemicals<br />
NICO P.E.VERMEULEN, RONALD T.H.van WELIE, BEN<br />
M.de ROOIJ and JAN N.M.COMMANDEUR<br />
Vrije Universiteit, Amsterdam<br />
Introduction<br />
In our industrialized world with increasing numbers <strong>of</strong> body foreign<br />
chemicals (xenobiotics) including drugs, food additives, pesticides,<br />
industrial chemicals and environmental pollutants, public concern about<br />
possible adverse (health) effects is growing. In 1989, for example, actual<br />
environmental topics in the Netherlands were photochemical summersmog<br />
and the presence <strong>of</strong> dioxines in milk <strong>of</strong> cows feeding in the neighbourhood<br />
<strong>of</strong> household refuse combustion furnaces and cable stills (CCRX, 1989). In<br />
this regard, most attention is paid to exposure to potentially mutagenic and<br />
carcinogenic xenobiotic chemicals. Apart from environmental exposure,<br />
especially at the workplace man may be exposed to elevated levels <strong>of</strong><br />
mixtures <strong>of</strong> known or unknown chemicals. Two centuries ago, cancer <strong>of</strong> the<br />
scrotum and testicles in chimney-sweepers was the first recognized<br />
occupational cancer (Pott, 1795). Since then numerous other hazardous<br />
occupational activities have been traced (Farmer et al., 1987).<br />
Nowadays, toxicologists are more and more focussed on the in vivo and<br />
in vitro bioactivation and bioinactivation mechanisms <strong>of</strong> chemicals. In the<br />
development <strong>of</strong> toxicity different stages are generally being distinguished:<br />
(1) toxicokinetics (absorption, distribution and elimination), (2)<br />
biotransformation, resulting in activation or inactivation <strong>of</strong> the chemicals,<br />
(3) reversible or irreversible interactions with cellular or tissue<br />
components, (4) protection and repair mechanisms and (5) nature and<br />
extent <strong>of</strong> the toxic effect for the organism (Vermeulen et al., 1990).<br />
Knowledge <strong>of</strong> for example species, dose, route <strong>of</strong> absorption, time <strong>of</strong><br />
exposure, tissue and organ selective interactions with (critical) cellular<br />
macro-molecules contributes to the understanding <strong>of</strong> molecular mechanisms<br />
<strong>of</strong> toxicity. Molecular mechanisms are useful in the prediction and<br />
prevention <strong>of</strong> chemically induced toxicities and they may play an<br />
important role in for example risk assessment and in the development <strong>of</strong><br />
safer chemicals (Vermeulen et al., 1990).
In this chapter, first the basic toxicokinetic concepts concerning the dis<br />
tribution, elimination and biotransformation <strong>of</strong> xenobiotics will be<br />
summarized. Subsequently, the relevance <strong>of</strong> these concepts will be<br />
illustrated and evaluated with the aid <strong>of</strong> a number <strong>of</strong> toxicokinetic studies<br />
in animals and humans concerning the nematocide 1,3-dichloropropene,<br />
the fungicide etridiazol, the chemical monomer 1,3-butadiene and the<br />
industrial solvent, 1,1,2-tri-chloroethylene. Apart from interspecies<br />
differences in the toxicokinetics, special attention will be given to<br />
interindividual differences in the toxicokinetics, among other things, as a<br />
result <strong>of</strong> genetically determined deficiencies in biotransformation enzymes<br />
as well as to its importance for the risk assessment <strong>of</strong> human exposure to<br />
industrial chemicals.<br />
Disposition <strong>of</strong> xenobiotics<br />
N.P.E.VERMEULEN ET AL. 13<br />
The overall fate <strong>of</strong> xenobiotics in an organism is determined by various<br />
toxicokinetic processes notably the route <strong>of</strong> administration, absorption,<br />
distribution and elimination. Chemicals may enter the body via various<br />
routes. Main routes are the lung, skin and gastrointestinal tract. The<br />
intraperitoneal, intramuscular, intravenous and subcutaneous routes are<br />
largely confined to experimental toxicological and therapeutic agents.<br />
Following absorption, xenobiotics enter the systemic or portal blood<br />
circulation. Distribution <strong>of</strong> chemicals in blood, organs and tissues usually<br />
occurs rapidly. The final plasma concentration depends on the ability <strong>of</strong><br />
the chemicals to pass cell membranes and on their affinity to various<br />
macromolecular proteins and tissues. Distribution to the kidney may result<br />
in direct excretion <strong>of</strong> the unchanged parent chemical. The physicochemical<br />
characteristics, such as lipophilicity and binding to plasma proteins, play an<br />
important role in the ultimate fate <strong>of</strong> a chemical in the body. The<br />
disposition <strong>of</strong> xenobiotics in the body is shown schematically in Figure 2.1.<br />
Its schematic relationship with biological/ toxicological effects is shown in<br />
Figure 2.2.<br />
Biotransformation plays an important role in the disposition <strong>of</strong><br />
xenobiotics in vivo. The liver is quantitatively the most important organ in<br />
the process <strong>of</strong> biotransformation. It receives a relative high bloodflow<br />
directly from the gastrointestinal tract via the portal vein, sometimes giving<br />
rise to the so-called hepatic ‘first-pass effect’ due to the presence <strong>of</strong> high<br />
concentrations <strong>of</strong> phase I and phase II metabolizing enzymes.<br />
Other important organs in biotransformation are the lungs, kidneys and<br />
the intestine. The primary object <strong>of</strong> biotransformation generally is to<br />
increase the hydrophilicity <strong>of</strong> chemicals, thus facilitating excretion by the<br />
kidneys in the urine or by the liver in the bile. Phase I reactions involve<br />
oxidation, reduction and hydrolysis reactions and phase II reactions<br />
conjugation or synthetic reactions. Phase I metabolic reactions generally
14 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Figure 2.1 Schematic representation <strong>of</strong> the fate <strong>of</strong> xenobiotics in the body according<br />
to their physico-chemical properties. Phase I and phase II represent the<br />
biotransformation processes. Adapted from Ariens and Simonis (1980).<br />
convert xenobiotic chemicals to more hydrophilic derivatives by<br />
introducing functional groups such as hydroxyl, sulphydryl and amino- or<br />
carboxylic acid groups. Phase II reactions are conjugation reactions in<br />
which the parent compounds or phase I derived metabolites are covalently<br />
bound to for example glucuronic acid, sulphate or glutathione.<br />
The group <strong>of</strong> cytochrome P-450 isoenzymes is the most important enzyme<br />
system in the catalysis <strong>of</strong> phase I reactions. The microsomal cytochrome<br />
P-450 system consists <strong>of</strong> various cytochrome P-450 isoenzymes and<br />
NADPH-cytochrome P450 reductase. It is involved in different metabolic<br />
reactions. At least three main types <strong>of</strong> activities can be distinguished,<br />
namely monooxygenase activity, oxidase activity and reductive activity<br />
(Guengerich 1994; Koymans et al., 1993). Glucuronic acid conjugation,<br />
catalyzed by UDP-glucuronyltransferases, represents one <strong>of</strong> the major<br />
phase II conjugation reactions in the conversion <strong>of</strong> exogenous and
endogenous chemicals. In mammals, another important conjugation<br />
reaction <strong>of</strong> hydroxyl groups is sulfatation, catalyzed by sulfotransferases<br />
(Sipes and Gandolfi, 1986). The group <strong>of</strong> glutathione S-transferase (GST)<br />
isoenzymes also represents an important phase II enzyme system. GST<br />
isoenzymes consist <strong>of</strong> two subunits on which the nomenclature is based<br />
(Warholm et al., 1986). The most important activity <strong>of</strong> GSTs is the<br />
catalysis <strong>of</strong> the conjugation <strong>of</strong> electrophilic, hydrophobic chemicals with<br />
the tripeptide glutathione (GSH). In general, GSH conjugation ultimately<br />
leads to the urinary excretion <strong>of</strong> mercapturic acids (N-acetyl-L-cysteine Sconjugates)<br />
(Vermeulen, 1989; Van Welie et al., 1992).<br />
Toxicokinetic principles<br />
General principles<br />
N.P.E.VERMEULEN ET AL. 15<br />
Figure 2.2 Disposition and biological effects <strong>of</strong> xenobiotics subdivided into three<br />
phases.<br />
The time course for the absorption, distribution, metabolism and<br />
elimination <strong>of</strong> a toxic substance is the subject <strong>of</strong> toxicokinetics. Implicit in<br />
any toxicokinetic description is the assumption that the response <strong>of</strong> target<br />
tissues or organs can be related to concentration pr<strong>of</strong>iles <strong>of</strong> the active form<br />
<strong>of</strong> the substance in that tissue or organ. Furthermore, it is <strong>of</strong>ten assumed<br />
that blood or plasma concentrations in one way or the other will reflect<br />
target tissue or organ concentrations and by inference the toxic effects.<br />
Under normal conditions one is generally dealing with first-order or linear<br />
kinetics, meaning that the amount <strong>of</strong> compound absorbed or eliminated<br />
(dQ) per unit <strong>of</strong> time (dt) is proportional to the total amount <strong>of</strong> compound<br />
present in the body. Zeroorder or non-linear kinetics may be valid as a<br />
consequence <strong>of</strong> various causes, e.g. saturation <strong>of</strong> binding <strong>of</strong> the toxic<br />
substance to plasma proteins or tissue components, or, more frequently
16 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Table 2.1 Frequently used toxicokinetic parameters and their formulas<br />
occurring, saturation <strong>of</strong> biotransformation enzyme systems. For the<br />
(mathematical) description <strong>of</strong> the toxicokinetics <strong>of</strong> substances, there exist<br />
at least two approaches at the moment: the traditional compartment<br />
pharmaco(toxico-)kinetic approach, in which the body is divided into one<br />
or more compartments, which do not necessarily correspond to<br />
physiological or anatomical units, and the physiologically-based pharmaco<br />
(toxico-)kinetic approach (PBPK or PBTK), in which organs, tissues and<br />
blood flow are taken into consideration. In Table 2.1 a summary <strong>of</strong> the most<br />
important and most frequently used traditional toxicokinetic parameters is<br />
shown. The value <strong>of</strong> some <strong>of</strong> these parameters is illustrated below, with the<br />
examples <strong>of</strong> 1,3-dichloropropene and etridiazol. The PBPK/PBTK approach<br />
is illustrated with the example <strong>of</strong> 1,3-butadiene.<br />
Principles <strong>of</strong> urinary excretion<br />
Of special interest in relation to this contribution also is the urinary<br />
excretion <strong>of</strong> xenobiotics and their metabolites by the kidneys. Two basic
N.P.E.VERMEULEN ET AL. 17<br />
processes, namely glomerular filtration and tubular secretion are used by<br />
the kidneys to remove chemicals from the bloodstream into the urine<br />
(Hook and Hewitt 1986). The kidneys are highly vulnerable to potential<br />
toxicants not only because they receive a high bloodflow (25% <strong>of</strong> the<br />
cardiac output), but also because they have the intrinsic ability to<br />
concentrate compounds. Recently, it has also become clear that xenobiotics<br />
may become nephrotoxic in the kidney itself due to bioactivation processes<br />
in combination with insufficient protection mechanisms (Commandeur and<br />
Vermeulen, 1991).<br />
The elimination <strong>of</strong> chemicals by the kidney is generally governed by firstorder<br />
processes. During first-order excretion kinetics the urinary<br />
elimination rate <strong>of</strong> a chemical is directly proportional to the plasma<br />
concentration. This means that the higher the plasma concentration the<br />
more <strong>of</strong> the chemical will be excreted in urine per unit <strong>of</strong> time. The urinary<br />
elimination rate (dQ/dt) can be calculated from a semi-logarithmic plot <strong>of</strong><br />
the urinary elimination rate versus the time <strong>of</strong> the intermittently collected<br />
urine samples (dQ/dt (mg h −l )=volume (1)×concentration (mg 1 −1 )/time (h))<br />
(Figure 2.3A).<br />
From the slope <strong>of</strong> the semi-logarithmic plasma concentration or urinary<br />
excretion rate versus time curve, the elimination rate constant (k el) and the<br />
urinary half-life <strong>of</strong> elimination (t 1/2) can be calculated. The half-life <strong>of</strong><br />
elimination is the time required to decrease the plasma concentration or the<br />
urinary elimination rate by one-half. The volume <strong>of</strong> distribution <strong>of</strong> the<br />
chemical normally can not be calculated from the urinary excretion data.<br />
Because the amount <strong>of</strong> chemical excreted in urine per unit <strong>of</strong> time (dQ/dt)<br />
is proportional to the plasma concentration (C p), the t 1/2 derived from the<br />
urinary elimination rate constant is identical to the t 1/2 <strong>of</strong> the chemical in<br />
plasma. It is evident that under these conditions the urinary excretion rate<br />
curve has the same shape as the plasma concentration curve (Figure 2.3B).<br />
In practice, the concentration <strong>of</strong> a chemical in urine (mg l −1 ) can be<br />
determined and multiplied by the volume (1) <strong>of</strong> the urine sample in order<br />
to calcu late the amount (mg) <strong>of</strong> chemical excreted over a period <strong>of</strong> time. In<br />
a semi-logarithmic plot the amount <strong>of</strong> chemical excreted is plotted against<br />
the midpoint <strong>of</strong> the interval <strong>of</strong> collection (Figure 2.3B). The accuracy <strong>of</strong> the<br />
method strongly depends on the way and the number <strong>of</strong> urine samples<br />
collected. As a rule <strong>of</strong> thumb, urine samples have to be collected during at<br />
least four half-lives <strong>of</strong> elimination. The complete cumulative urinary<br />
excretion <strong>of</strong> a chemical can be calculated as the area under the urinary<br />
excretion rate versus time curve including extrapolation time to infinity.<br />
Occupational exposure to chemicals frequently occurs 5 days a week, 8 h<br />
a day, with an exposure free period <strong>of</strong> 16 h. Intermittent exposure to a<br />
chemical may lead to different accumulation situations in the body<br />
depending on the periods between exposure in relation to t 1/2 (Table 2.1).<br />
No accumulation will occur when the intervals between the exposure
18 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Figure 2.3 Schematic representation <strong>of</strong> first order kinetics <strong>of</strong> (A) the plasma<br />
concentration (C p) <strong>of</strong> a chemical versus the urinary elimination rate (dQ/dt), (B) the<br />
relation between the elimination rate in plasma and urine and (C) the cumulative<br />
excretion ( (%)) versus time. In (B): slope=–k el/2.303 and t 1/2=0.693/k el.<br />
Figure 2.4 Urinary excretion <strong>of</strong> a hypothetical metabolite during 3 days <strong>of</strong><br />
intermittent exposure: t 1/2
urinary concentration at certain time points the net total cumulative<br />
excretion <strong>of</strong> day 2 can be calculated.<br />
Monitoring in occupational toxicology<br />
N.P.E.VERMEULEN ET AL. 19<br />
In occupational toxicology generally four monitoring approaches are<br />
distinguished, namely: environmental monitoring (EM), biological<br />
monitoring (BM), biological effect monitoring (BEM) and health<br />
surveillance (HS) (Figure 2.5). EM and BM are concerned with the<br />
measurement and assessment <strong>of</strong> ambient exposure and health risk<br />
compared to appropriate references. EM determines xenobiotics at the<br />
workplace, BM determines xenobiotics or their metabolites in tissues or<br />
secreta. BEM is concerned with the measurement and assessment <strong>of</strong> early,<br />
non-adverse, biological alterations in exposed workers to evaluate<br />
exposure and/or health risk compared to appropriate references. HS is<br />
concerned with periodic medico-physiological examination <strong>of</strong> exposed<br />
workers with the objective <strong>of</strong> protecting and preventing occupationally<br />
related diseases (Zielhuis and Henderson, 1986).<br />
EM was shown to be <strong>of</strong> limited value for assessing the internal dose <strong>of</strong> a<br />
chemical by not taking into account for example toxicokinetic and<br />
toxicodynamic processes determining the ultimate fate <strong>of</strong> xenobiotics in the<br />
body. To a certain extent, BM appeared to overcome the problems<br />
inherently related to EM. BM assesses the overall exposure to xenobiotics<br />
that are present at the workplace through measurement <strong>of</strong> the appropriate<br />
determinant(s) in biological specimens collected from the worker at specific<br />
timepoints (ACGIH, 1990).<br />
Ideally, not only the relation between exposure and effect is known, but<br />
also the toxicokinetic and toxicodynamic interactions linking these two. If<br />
these processes are elucidated, quantitative knowledge <strong>of</strong> a determinant <strong>of</strong><br />
one <strong>of</strong> the different monitoring methods allows an assessment either <strong>of</strong> the<br />
level <strong>of</strong> exposure or <strong>of</strong> the level <strong>of</strong> effect (Figure 2.5). For example, the<br />
level <strong>of</strong> urinary mercapturic acid excretion could assess the potential health<br />
hazard <strong>of</strong> an occupational exposure situation (Henderson et al., 1989).<br />
In practice, a complete view on the relation between toxicokinetics and<br />
toxicodynamics has not been elucidated for a single chemical up to now.<br />
Occupational monitoring methods all have their specific values based on<br />
their selectivity, sensitivity, validity and logistics and should therefore be<br />
used complementary to each other. All methods operate on the continuum<br />
from exposure to effect, the limits between which occupational toxicology<br />
studies operate.
20 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Figure 2.5 Occupational monitoring methods and their relation to exposure versus<br />
effect assessment and to toxicokinetic and toxicodynamic processes. (Adapted from<br />
Henderson et al., 1989).<br />
Glutathione conjugation products as biomarkers<br />
In principle, GSH-conjugation derived metabolites can be used as a<br />
biomarker <strong>of</strong> internal dose. Glutathione (GSH), a tripeptide consisting <strong>of</strong><br />
the amino acids glycine, cysteine and -glutamine, plays an important role<br />
in the detoxification <strong>of</strong> potentially electrophilic chemicals or metabolites. In<br />
contrast, toxification via GSH-conjugation, for example <strong>of</strong> 1,2dibromoethane,<br />
hexachlorobutadiene, benzyl- and allylisothiocyanate has<br />
also been reported. β-lyase dependent bioactivation <strong>of</strong> cysteine-conjugates,<br />
derived from the initially formed GSH-conjugates, sometimes resulted in<br />
the formation <strong>of</strong> new reactive intermediates which are responsible for<br />
carcinogenic, mutagenic and other toxicological effects (Vermeulen, 1989;<br />
Van Welie et al., 1992).<br />
The initial step in GSH-conjugation is reaction <strong>of</strong> the nucleophilic<br />
sulphhydryl with electrophilic centers <strong>of</strong> a chemical. GSH-conjugation is<br />
catalysed by a family <strong>of</strong> glutathione S-transferase (GST) enzymes. A wide<br />
range <strong>of</strong> chemicals can be handled by this enzyme system due to the
existence <strong>of</strong> a large number <strong>of</strong> isoenzymes with different, though<br />
overlapping, substrate selectivity. The final detoxification capacity through<br />
GSH and GST enzymes <strong>of</strong> an organism depends on endogenous factors<br />
such as tissue distribution, genetic deficiencies, aging and hormonal<br />
influences and on exogenous factors such as sensitivity to inhibition and<br />
induction <strong>of</strong> GSTs (Vermeulen, 1989; Van Welie et al., 1992).<br />
GSH-conjugates normally are not excreted unchanged in urine or faeces.<br />
Catabolism <strong>of</strong> the GSH-conjugates results in the formation and excretion<br />
<strong>of</strong> a variety <strong>of</strong> sulphur containing metabolites, among which thioethers and<br />
mercapturic acids (S-substituted N-acetyl-cysteine conjugates) belong to the<br />
most important. The mercapturic acid pathway is shown in Figure 2.6.<br />
Thioethers in human studies<br />
N.P.E.VERMEULEN ET AL. 21<br />
Figure 2.6 Schematic representation <strong>of</strong> the mercapturic acid pathway: GSHconjugation<br />
with an electrophilic chemical (RX) and the biosynthesis to a<br />
mercapturic acid. E1: glutathione S-transferase, E2: -glutamyltranspeptidase, E3:<br />
cysteinylglycinase and aminopeptidase, E4: cysteine conjugate N-acetyltransferase,<br />
E5: N-deacetylase.<br />
Several years ago, Seutter-Berlage et al. proposed the appearance <strong>of</strong><br />
thioethers such as mercapturic acids (R-S-R′), mercaptans (R-SH) and<br />
disulfides (R-S-S-R′) in urine as an indicator <strong>of</strong> exposure to potentially<br />
alkylating chemicals. The thioether assay is an aselective assay to detect<br />
metabolic end-products excreted in urine <strong>of</strong> (non)occupational exposure to<br />
various electrophilic chemicals. It includes three steps, namely: (i)<br />
extraction, (ii) alkaline hydrolysis and (iii) derivatization, subsequently<br />
followed by spectrophotometric analysis at 412 nm. The thioether assay
22 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Figure 2.7 Urinary excretion <strong>of</strong> thioethers (mmol SH/mol creatinine), <strong>of</strong> applicators<br />
exposed to 3.8 (− −), 9.8 (− −) and 18.9 (− −) mg m −3 8-h TWA (Z+E)-1,3dichloropropene<br />
in respiratory air, respectively. Darker shaded areas indicate<br />
exposure periods.<br />
was first applied to compare thioether excretions in urine <strong>of</strong> employees <strong>of</strong> a<br />
chemical plant. Highest thioether excretions were found in rubber workers<br />
and radial tyre builders when compared with clerks, plastic monomer<br />
mixers and footwear preparers. Recently, urinary thioether excretion was<br />
related to the occupational respiratory exposure <strong>of</strong> applicators in the Dutch<br />
flower-bulb culture to 1,3-dichloropropene (DCP) (Van Welie et al.,<br />
1991a). Instead <strong>of</strong> a discrete comparison <strong>of</strong> thioether excretion with<br />
exposed versus non-exposed groups, in this study thioether excretion was<br />
related to a continuous scale <strong>of</strong> airborne DCP concentrations.<br />
Significant linear relations between respiratory exposure to DCP and postminus<br />
preshift thioether concentration and cumulative thioether excretion<br />
were found. The urinary excretion <strong>of</strong> DCP-thioethers followed first-order<br />
elimination kinetics (Figure 2.7) with half-lives <strong>of</strong> elimination <strong>of</strong> 8.0±2.5 h<br />
(n=5) based on urinary excretion rates and 9.5±3.1 h (n=5) based on<br />
creatinine excretion. The elimination half-lives <strong>of</strong> the thioethers were<br />
almost two fold higher when compared to the half-lives <strong>of</strong> elimination <strong>of</strong><br />
the mercapturic acids <strong>of</strong> Z-and E-1,3 dichloropropene. This illustrates the<br />
main problem <strong>of</strong> urinary thioethers, viz. high background levels originating<br />
from endogenous or exogenous sources, such as smoking and diet (e.g.<br />
horse radish, onion and garlic).
Mercapturic acids in human studies<br />
Mercapturic acids, S-substituted N-acetyl-L-cysteine S-conjugates, in urine<br />
can be used as biomarkers <strong>of</strong> internal dose <strong>of</strong> electrophilic xenobiotics.<br />
Mercapturic acids are metabolic end products <strong>of</strong> GSH-conjugation <strong>of</strong><br />
various potentially electrophilic chemicals (Figure 2.6). The first<br />
mercapturic acids were identified in 1879 as sulphur containing<br />
metabolites after administration <strong>of</strong> bromobenzene to dogs (see references in<br />
Vermeulen, 1989). Since then mercapturic acids from many chemicals have<br />
been identified and these types <strong>of</strong> urinary metabolites have been used in<br />
biotransformation, biological monitoring and toxicological studies<br />
(Vermeulen, 1989; Van Welie et al., 1992).<br />
Commercial availability <strong>of</strong> reference compounds and the development <strong>of</strong><br />
a number <strong>of</strong> different analytical techniques attributed to the popularity <strong>of</strong><br />
mercapturic acids in biological monitoring studies during the last few<br />
years. Urinary excretion <strong>of</strong> the stereoisomeric mercapturic acids <strong>of</strong> Z- and<br />
E-1,3-dichloropropene, a soil fumigant frequently used in agriculture,<br />
proved to be a suitable biomarker for the exposure to both isomers in man.<br />
Strong correlations were observed between 8-h time weighted average<br />
exposure to Z- and E-DCP and complete cumulative excretion <strong>of</strong> N-acetyl-<br />
S-(Z- and E-3-chloropropenyl-2)-L-cysteine in urine. N-acetyl-S-<br />
(cyanoethyl)-L-cysteine was proposed as biomarker <strong>of</strong> exposure to<br />
acrylonitrile. The best correlation between uptake <strong>of</strong> acrylonitrile via the<br />
lungs and excretion <strong>of</strong> the cyanoethyl mercapturic acid in urine was<br />
obtained in samples collected between the sixth and the eighth hour after<br />
the beginning <strong>of</strong> exposure (Jakubowoski et al., 1987). The phenyl<br />
mercapturic acid <strong>of</strong> benzene was regarded as a useful biomarker <strong>of</strong><br />
exposure below 1 ppm <strong>of</strong> workers in a chemical production plant<br />
(Stommel et al., 1989). The use <strong>of</strong> certain foodstuffs and drugs may also<br />
give rise to the excretion <strong>of</strong> mercapturic acids. Consumption <strong>of</strong> cabbage<br />
and horse radish for example gave rise to increased thioether excretion.<br />
Consumption <strong>of</strong> garlic and onions resulted in the excretion <strong>of</strong> N-acetyl-S-<br />
(allyl- and 2-carboxypropyl)-L-cysteine in urine (Van Welie et al., 1992).<br />
The hypnotic drug ( α-bromo-isovalerylurea<br />
also gave rise to the excretion<br />
<strong>of</strong> two diastereomeric α-bromoisovalerylurea<br />
mercapturic acid conjugates<br />
in urine (Mulders et al., 1993). S-Phenyl mercapturic acid was present in<br />
urine <strong>of</strong> groups <strong>of</strong> smokers and non-smokers, not exposed to benzene, in<br />
concentrations <strong>of</strong> 4.0±4.0 µg g −1 creatinine (Stommel et al., 1989).<br />
Toxicokinetics<br />
N.P.E.VERMEULEN ET AL. 23<br />
Knowledge about the toxicokinetics <strong>of</strong> mercapturic acids is necessary to<br />
develop optimal sampling strategies in occupational studies. Urinary<br />
excretion rates <strong>of</strong> mercapturic acids theoretically may reflect the rates <strong>of</strong>
24 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Figure 2.8 Urinary excretion (– –=Z, − −=E) and cumulative excretion (– –=Z,<br />
− −=E) <strong>of</strong> Z- and E-DCP-MA <strong>of</strong> an applicator due to an 8-h TWA respiratory<br />
exposure to 2.32 mg m −3 Z-DCP and 1.73 mg m −3 E-DCP. In (A) the mercapturic acid<br />
excretion rate is depicted and in (B) the mercapturic acid excretion based on<br />
creatinine excretion.<br />
elimination <strong>of</strong> the parent compounds from blood and can be used to<br />
calculate the (complete) cumulative excretion <strong>of</strong> mercapturic acids related<br />
to exposure. By knowing an individual’s mercapturic acid excretion rate,<br />
the contribution to urinary mercapturic acid excretion <strong>of</strong> the day under<br />
study on succeeding day(s) can be calculated. The contributions <strong>of</strong> previous<br />
days <strong>of</strong> exposure can also be used to correct the mercapturic acid excretion<br />
<strong>of</strong> the exposure day under study. The urinary half-life <strong>of</strong> elimination is<br />
inversely proportional to the elimination rate constant. The urinary halflife<br />
<strong>of</strong> both mercapturic acids <strong>of</strong> Z- and E-DCP in man was ca. 5 h<br />
(Figure 2.8) and they were not significantly different, i.e. 5.0±1.2 h for Z-<br />
DCP-MA and 4.7±1.3 h for E-DCP-MA. Strong corre-lations (r≥0.93) were<br />
observed between respiratory 8 h time weighted average (TWA) exposure<br />
to Z- and E-DCP and complete cumulative urinary excretion <strong>of</strong> Z- and E-<br />
DCP-MA. There is still a lack <strong>of</strong> knowledge about the magnitude <strong>of</strong> the
N.P.E.VERMEULEN ET AL. 25<br />
intra- and inter-individual differences in GSH-conjugation and mercapturic<br />
acid excretion. Factors causing these differences are sex, stress, diet, age,<br />
enzyme induction and inhibition, pathology and genetic variability. Apart<br />
from these factors the presence or absence <strong>of</strong> glutathione S-transferases<br />
(GSTs) or GST activity in different persons is <strong>of</strong> special interest in relation<br />
to urinary mercapturic acid excretion. The most intriguing factor known in<br />
this context is the human genetic polymorphism <strong>of</strong> mu-class GSTs. The<br />
GST isoenzyme µ is expressed only in approximately 60% <strong>of</strong> the human<br />
population. Mu-class GST isoenzymes showed a high specific activity<br />
towards for example styrene-7,8-oxide and benzo(a)pyrene-4,5dihydrodiol-4,5-oxide<br />
and E- and Z-DCP. Genetic polymorphism <strong>of</strong> muclass<br />
GSTs was postulated as a determinant in the excretion <strong>of</strong> the<br />
mercapturic acids <strong>of</strong> Z- and E-DCP in occupationally exposed applicators.<br />
However, between mu-class positive (n=9) and mu-class<br />
Table 2.2 Urinary excretion levels, urinary ratios and half-lives <strong>of</strong> elimination <strong>of</strong> Zand<br />
E-DCP mercapturic acids <strong>of</strong> mu-class positive and mu-class negative<br />
individuals a<br />
a Urinary excretion level represents the cumulative excretion <strong>of</strong> Z- and E-DCP-MA<br />
in 0–36 h urine, corrected for the time weighted average 8-h exposure to Z- and E-<br />
DCP. Values are expressed as means±SD for the number <strong>of</strong> individuals indicated in<br />
parentheses.<br />
b (mmol mercapturic acid)/(mmol DCP m −3 ).<br />
c Z-DCP-MA/E-DCP-MA<br />
d Half-life <strong>of</strong> elimination<br />
negative (n=3) applicators, neither a difference in urinary half-lives <strong>of</strong><br />
elimination nor in cumulative excretion <strong>of</strong> both mercapturic acids <strong>of</strong> Zand<br />
E-DCP was seen (Vos et al., 1991) (Table 2.2).<br />
α-Bromoisovalerylurea,<br />
a sedative and hypnotic drug, is a racemic drug<br />
which is also metabolized by GSH-conjugation. It was proposed as a<br />
model substrate to study the pharmacokinetics and stereoselectivity <strong>of</strong> GSHconjugation<br />
in humans. Stereoselective mercapturic acid formation <strong>of</strong> Rand<br />
S-α-bromoisovalerylurea was seen in in vitro studies with purified GST<br />
isoenzymes and in vivo in rat and man. In humans, a pronounced<br />
stereoselectivity in urinary mercapturic acid excretion was observed. Of an<br />
oral dose <strong>of</strong> R- and S-α-bromoisovalerylurea, 22.5±4.3 and 5.7±1.6% was<br />
excreted as mercapturic acid in 24 h, respectively. The half-lives <strong>of</strong><br />
elimination <strong>of</strong> both diastereoisomeric mercapturic acids were 1.5±0.4 and<br />
3.1±1.3 h, respectively. Both the pharmacokinetics <strong>of</strong> α-bromoisovaleryl
26 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Figure 2.9 Proposed biotransformation pathway <strong>of</strong> etridiazol leading to 5-ethoxy-1,<br />
2,4-thiadiazole-3-carboxylic acid (ET-CA) and N-acetyl-S-(ethoxy-1,2,4thiadiazol-3-yl-methyl)-L-cysteine<br />
(ET-MA) in rat and humans. Unidentified<br />
intermediates are presented between brackets ([...]). GSH: glutathione, MAP:<br />
mercapturic acid pathway.<br />
ureas and their stereoselectivity, however, were not found to be different for<br />
subjects who were GSH S-transferase class mu deficient and subjects who<br />
were not (Mulders et al., 1993).<br />
Disposition <strong>of</strong> etridiazol<br />
Etridiazol (Aaterra; 5-ethoxy-3-trichloromethyl-l,2,4-thiadiazole<br />
(Figure 2.9)) is an agricultural fungicide used to control phycomycetous<br />
fungi in, for example, plants, tomatoes, cucumbers, cauliflowers and<br />
celery. Concerning external exposure <strong>of</strong> applicators (e.g. greenhouse<br />
handgunners and foggers) it has been concluded that exposure may occur<br />
through inhalation and dermal absorption. For the purpose <strong>of</strong> the<br />
development <strong>of</strong> a biomonitoring assay disposition studies were performed<br />
recently in rats and human volunteers (Van Welie et al., 1991c). Two<br />
metabolites, 5-ethoxy-l,2,4-thiadiazole-3-carboxylic acid (ET-CA) and a<br />
mercapturic acid, N-acetyl-S-(5-ethoxy-l,2,4-thiadiazol-3-yl-methyl)-Lcysteine<br />
(ET-MA) were identified as new metabolites. Based on a<br />
preliminary toxicokinetic study, the urinary excretion <strong>of</strong> the former<br />
metabolite amounted to 22±9% <strong>of</strong> an oral dose <strong>of</strong> etridiazol (while ET-MA<br />
and unchanged etridiazol were less than 1 % <strong>of</strong> the dose), ET-CA was<br />
proposed as a possible biomarker <strong>of</strong> exposure to this fungicide.
1,1,2-Trichloroethylene<br />
N.P.E.VERMEULEN ET AL. 27<br />
The solvent properties <strong>of</strong> 1,1,2-trichloroethylene (TRI) have resulted in its<br />
widespread use in metal degreasing and a wide variety <strong>of</strong> other industrial<br />
applications. TRI has now been in common use for more than 50 years.<br />
During this period <strong>of</strong> time, workers have been exposed to a wide range <strong>of</strong><br />
concentrations, in some cases for periods <strong>of</strong> 25 years or longer. This has<br />
allowed the compilation <strong>of</strong> a great data base about the effects <strong>of</strong> TRI on<br />
human health. Moreover, information has been supplemented by<br />
numerous studies in experimental animals.<br />
Epidemiological studies on more than 15000 individuals with a followup<br />
<strong>of</strong> more than 25 years have shown no evidence <strong>of</strong> an association<br />
between human exposure to TRI and increased incidence <strong>of</strong> cancer or<br />
cancer mortality. However, several <strong>of</strong> these studies had more or less serious<br />
shortcomings. A summary <strong>of</strong> effects related to TRI and/or TRI-related<br />
metabolism is given in Table 2.3. These and other data are taken from<br />
Goeptar et al., 1995a.<br />
An increased incidence <strong>of</strong> lung tumors has been reported in female<br />
B 6C 3F 1 and male Swiss mice exposed to TRI by inhalation. The effect was<br />
not observed in male B 6C 3F 1 nor in female Swiss mice nor in rats. This<br />
apparent strain-, sex- and lung-specific response fails to resolve the issue <strong>of</strong><br />
whether or not TRI is a carcinogenic hazard to man. Mechanistic studies<br />
on mouse lung tumor formation have explained the sex and species<br />
differences. In this context, chloral formation (Figure 2.10) in Clara cells,<br />
containing relatively high cytochrome P-450 concentrations, has been<br />
identified to be responsible for the development <strong>of</strong> mouse lung tumors.<br />
Importantly, lung tumors have not been found in humans after long-term<br />
occupational exposure in TRI.<br />
TRI causes an increase in the incidence <strong>of</strong> liver cancer in both sexes <strong>of</strong><br />
B 6C 3F 1 and Swiss mice following either gavage or inhalatory exposure, but<br />
not in NMRI and Ha: ICR mice nor in rats. A rodent specific link between<br />
peroxisome proliferation, DNA synthesis, inhibition <strong>of</strong> intercellular<br />
communication and cancer (Table 2.3) suggests that these responses are the<br />
basis <strong>of</strong> the hepatocarcinogenicity induced by TRI. The identification <strong>of</strong><br />
TCA in cancer bioassays as the responsible metabolite for these effects<br />
confirmed this hypothesis. However, when TCA was administered to both<br />
rats and mice, liver cancer was only observed in mice and not in rats. The<br />
reason for this species selectivity in liver effects is explained by the kinetic<br />
behavior <strong>of</strong> TRI and TCA in rodents. Both rats and mice have a considerable<br />
capacity to metabolize TRI to TCA and TCE, the maximal capacities being<br />
closely related to the relative surface areas rather than to their body<br />
weights. Oxidative metabolism <strong>of</strong> TRI in rats is linearly related to dose at<br />
lower dose levels, but it becomes saturated at higher dose levels. Thus, an<br />
important difference between rats and mice is the lower saturation
28 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Table 2.3 Reported toxic effects related to TRI and/or TRI-derived metabolites<br />
a References: see review Goeptar et al., 1995b.<br />
n.d.: not determined.
N.P.E.VERMEULEN ET AL. 29<br />
Figure 2.10 Oxidative metabolism <strong>of</strong> TRI in the rodent and mammalian liver and<br />
the formation <strong>of</strong> metabolites which are excreted in the urine.<br />
concentration in the former species. The relevance <strong>of</strong> the mechanisms <strong>of</strong><br />
liver tumor formation in B 6C 3F 1 and Swiss mice for humans exposed to<br />
TRI has been assessed in studies comparing metabolic rates in mice, rats<br />
and humans. In contrast to the rat, the oxidative metabolism <strong>of</strong> TRI to<br />
TCA in humans is not limited by saturation. In this respect, humans<br />
resemble the mouse and might be able to produce sufficient TCA to induce<br />
peroxisome proliferation and consequently liver cancer. However, there are<br />
significant differences between mice and humans. First, humans metabolize<br />
approximately 60 times less TRI on a body weight basis than mice at<br />
similar exposure levels. Second, TCA has been shown to induce<br />
peroxisome proliferation in mouse hepatocytes but not in human<br />
hepatocytes (Table 2.3). Consequently, the combination <strong>of</strong> extensive<br />
oxidative metabolism <strong>of</strong> TRI to TCA and the ability <strong>of</strong> TCA to induce<br />
peroxisome proliferation appear to be unique to B 6C 3F 1 and Swiss mice.<br />
TRI-induced renal toxicity and tumors were found in Sprague-Dawley,<br />
Fischer 344 and Osborne-Mendel rats. These nephrocarcinogenic effects <strong>of</strong><br />
TRI were specific to male rats and were not seen in female rats nor in mice<br />
<strong>of</strong> either sex. 1,2-DCV-Cys, formed from TRI via the mercapturic acid<br />
pathway, has been identified as a likely metabolite involved in the observed<br />
renal toxicity and probably also in renal carcinogenicity in rats. TRI is<br />
metabolized by a minor pathway involving initial hepatic GSH-conjugation<br />
<strong>of</strong> TRI. The resulting DCV-G is further metabolized (Figure 2.11) and<br />
excreted in urine as two regioisomeric mercapturic acids, namely vicinal 1,<br />
2-DCV-Nac and geminal 2,2-DCV-Nac (Figure 2.11). 1,2-DCV-Cys (the<br />
precursors <strong>of</strong> 1,2-DCV-Nac) is a substrate for the renal L-cysteine S-
30 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Figure 2.11 Possible routes <strong>of</strong> metabolism <strong>of</strong> S-(1,2-dichlorovinyl)glutathione (1,2-<br />
DCV-G). Steps are catalyzed by (a) -glutamyltransferase; (b) cysteinylglycine<br />
dipeptidase; (c) L-cysteine S-conjugate β-lyase; (d) L-cysteine S-conjugate Nacetyltransferase;<br />
(e) acylase.<br />
conjugate β-lyase and it is more mutagenic and cytotoxic than 2,2-DCV-<br />
Cys (the precursors <strong>of</strong> 2,2-DCV-Nac).<br />
The bioactivation <strong>of</strong> 1,2-DCV-Cys is without a doubt a crucial step in<br />
the onset <strong>of</strong> nephrotoxicity in the rat, although the precise biological<br />
mechanisms by which these metabolites exert their nephrocarcinogenic<br />
effects are not yet fully understood. A key aspect in the onset <strong>of</strong><br />
nephrocarcinogenicity in rats, however, is that it will not occur in the<br />
absence <strong>of</strong> nephrotoxicity. This suggests that the alkylating effects <strong>of</strong> the<br />
reactive metabolites (most likely thioketenes) derived from bioactivation <strong>of</strong><br />
1,2-DCV-Cys by β-lyase may not be sufficient to cause kidney tumors. The<br />
specific activity <strong>of</strong> β-lyase, the key enzyme involved in the bioactivation <strong>of</strong><br />
DCV-Cys isomers, is similar in humans to that in the mouse and only 10%<br />
<strong>of</strong> that in the rat. Moreover, human TRI metabolism via the mercapturic<br />
acid pathway resembles that <strong>of</strong> the mouse. It is, therefore, questionable<br />
whether humans are able to produce sufficient DCV-Cys isomers from TRI<br />
to cause first nephrotoxicity and then nephrocarcinogenicity. An important<br />
finding is also that the occurrence <strong>of</strong> nephrotoxicity and
N.P.E.VERMEULEN ET AL. 31<br />
nephrocarcinogenicity in the male rat is dose-dependent. More specifically,<br />
cytotoxic kidney damage is a feature <strong>of</strong> high continuous exposure to TRI<br />
over prolonged periods <strong>of</strong> time. This is unlikely to occur in humans during<br />
occupational exposure. In fact, TRI has been found not to be nephrotoxic<br />
in humans chronically exposed to low levels <strong>of</strong> TRI (50 mg m −3 ).<br />
Consequently, it is unlikely that the renal tumors which are seen in rats at<br />
nephrotoxic dose levels <strong>of</strong> TRI and which are related to β-lyase mediated<br />
bioactivation <strong>of</strong> 1,2-DCV-Cys, are relevant to human health hazards at<br />
reasonably foreseeable levels <strong>of</strong> exposure.<br />
Physiologically based toxicokinetic modeling <strong>of</strong> 1,3butadiene<br />
Physiologically based pharmaco(toxico)-kinetic models differ from the<br />
conventional compartmental models in that they are based to a large extent<br />
on the actual physiology <strong>of</strong> the organism. Instead <strong>of</strong> compartments defined<br />
largely by the experimental data themselves, actual organ and tissue groups<br />
are used with weights and blood flows from the literature (Bisch<strong>of</strong> and<br />
Brown, 1966). Instead <strong>of</strong> composite rate constants determined by fitting<br />
the actual experimental data, physical-chemical and biochemical constants<br />
<strong>of</strong> the compound are used. The result is a mode which predicts the<br />
qualitative behavior <strong>of</strong> the experimental time course without being based<br />
on it. Refinements <strong>of</strong> the model to incorporate additional insights gained<br />
from comparison with experimental data yields a model which can be used<br />
for quantitative extrapolations well beyond the range <strong>of</strong> experiments. In<br />
recent years several PBTK- and PBPK-models have been published: for<br />
methylene chloride, see Andersen et al., 1987; for a review see Leung et al.,<br />
1988; for 1,3-butadiene, see Evelo et al., 1993.<br />
The development <strong>of</strong> a PBTK/PBPK model can be divided into a number<br />
<strong>of</strong> steps: (a) inventory <strong>of</strong> physiological and toxicological behaviour <strong>of</strong> the<br />
compound, (b) mathematical description <strong>of</strong> the biochemical/(patho)<br />
physiological processes involved, (c) parameterization <strong>of</strong> the mathematical<br />
descriptions, (d) the construction <strong>of</strong> the model, (e) refinement and<br />
validation <strong>of</strong> the model and (f) use <strong>of</strong> the predictions and risk assessment.<br />
As an illustrative example <strong>of</strong> this approach the recently described PBTKmodeling<br />
<strong>of</strong> 1,3-butadiene disposition and toxicity might be used (Evelo et<br />
al., 1993). 1,3-Butadiene used for the production <strong>of</strong> styrene-butadiene<br />
rubber, is known amongst others to cause lung carcinogenicity. In the rat<br />
the carcinogenicity <strong>of</strong> 1,3-butadiene is less pronounced while the evidence<br />
for human carcinogenicity is inconclusive, Monoand di-epoxy-butadiene<br />
are reactive metabolites held responsible for this effect. Butadiene<br />
monoxide is formed by microsomal fractions <strong>of</strong> the lung and liver <strong>of</strong> several
32 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS<br />
Figure 2.12 Physiologically based toxicokinetic model for description <strong>of</strong> butadiene<br />
distribution and metabolism in mice, rats and humans. Gas exchange occurs in the<br />
alveoli <strong>of</strong> the lung. Metabolism occurs in both the alveolar and bronchial areas <strong>of</strong><br />
the lung and in the liver. Metabolic activity in the three other compartments is<br />
ignored (Evelo et al., 1993).<br />
species. There are, however, large interspecies differences in the lung vs<br />
liver activities: mice>rats>humans/monkeys.<br />
The PBTK model used to describe butadiene distribution and metabolism<br />
in mice, rats and humans is shown in Figure 2.12. Gas exchange is<br />
supposed to occur in the alveoli <strong>of</strong> the lung and metabolism in both the<br />
alveolar and bronchial areas <strong>of</strong> the lung and in the liver. By using the<br />
experimentally determined or estimated species-selective parameters for
volumes, masses and blood flows <strong>of</strong> different organs, partition coefficients<br />
<strong>of</strong> 1,3-butadiene between blood and organs/tissues and for metabolic<br />
capacities in liver and lung (bronchial and alveolar areas), accurate dosedependent<br />
simulations were performed for the uptake <strong>of</strong> 1,3-butadiene in<br />
mice and rats in gas-closed chambers. Moreover, with the resulting model<br />
the relative importance <strong>of</strong> lung metabolism as compared to metabolism in<br />
the liver was predicted for the three different species. Lung metabolism<br />
appeared to be much more important than liver metabolism in mice, this in<br />
contrast to the situation in the rat and humans. Moreover, at low exposure<br />
concentrations the relative importance <strong>of</strong> lung metabolism was predicted to<br />
increase in mice as a result <strong>of</strong> diminished saturation <strong>of</strong> metabolism in this<br />
species. It was concluded that the observed species differences in lung vs<br />
liver metabolism <strong>of</strong> 1,3-butadiene (mice>rat>human) and the tendency<br />
towards increased lung metabolism at low doses might rationalize the<br />
observed species differences in the lung carcinogenicity <strong>of</strong> 1,3-butadiene<br />
and this knowledge should be useful in the in vivo extrapolation from high<br />
dose to low dose risk assessments within one species as well as in<br />
interspecies risk assessment extrapolations.<br />
Conclusions<br />
In conclusion, a pr<strong>of</strong>ound knowledge <strong>of</strong> the biodisposition and the<br />
toxicokinetics <strong>of</strong> a toxic or potentially toxic chemical is <strong>of</strong> utmost<br />
importance to the design and interpretation <strong>of</strong> laboratory assessments <strong>of</strong><br />
toxicity, to explain interspecies differences in toxicities and to extrapolate<br />
more reliably from animal experiments to man in the process <strong>of</strong> risk<br />
assessment. This also holds true for the design for proper biological<br />
monitoring procedures and for the interpretation <strong>of</strong> the results in terms <strong>of</strong><br />
potential health risks <strong>of</strong> exposure to chemicals. Apart from traditional<br />
compartment-based toxicokinetic approaches, more recent physiologicallybased<br />
toxicokinetics modeling approaches have distinct advantages for the<br />
above-mentioned purposes.<br />
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N.P.E.VERMEULEN ET AL. 33<br />
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GOEPTAR, A.R., SCHEERENS, H. and VERMEULEN, N.P.E., 1995b, Oxygen<br />
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JAKUBOWOSKI, M., LINHART, I., PIELAS, G. and KOPECKY, J., 1987, 2-<br />
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KOYMANS, L., DONNÉ-OP DEN KELDER, G.M., TE KOPPELE, J.M. and<br />
VERMEULEN, N.P.E., 1993, Cytochromes P450: their active-site structure<br />
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LEUNG, H.W., Ku, R.H., PAUSTENBACH, D.J. and ANDERSEN, M.E., 1988, A<br />
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MULDERS, T.M.T., VENIZELOS, V., SCHOEMAKER, R., COHEN, A.F.,<br />
BREIMER, D.D. and MULDER, G.J., 1993, Characterization <strong>of</strong> glutathione<br />
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P.T.H., 1979, Urinary mercapturic acid excretion as a biological parameter <strong>of</strong><br />
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STOMMEL, P., MÜLLER, G., STÜCKER, W., VERKOYEN, C., SCHÖBEL, S.<br />
and NORPOTH, K., 1989, Determination <strong>of</strong> S-phenylmercapturic acid in the<br />
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VAN WELIE, R.T.H., VAN MARREWIJK C.M., DE WOLFF, F.A. and<br />
VERMEULEN, N.P.E., 1991a, Thioether excretion in urine <strong>of</strong> applicators<br />
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3<br />
Metabolic Activation <strong>of</strong> <strong>Industrial</strong> Chemicals and<br />
Implications for Toxicity<br />
GERARD J.MULDER<br />
Leiden University, Leiden<br />
Introduction<br />
In the toxicity <strong>of</strong> industrial chemicals bioactivation (Anders, 1985) plays an<br />
important role. Obviously, its importance depends on the structure <strong>of</strong> the<br />
chemical as well as the toxic effect considered. Thus, inorganic compounds<br />
in general will not require bioactivation: metal salts or oxides will usually<br />
cause toxicity in the form in which they are taken up. However, even these<br />
chemicals may require further metabolism for maximum toxicity in the<br />
body: inorganic mercury may be converted to an organic form<br />
(methylmercury), and nitrate may be reduced to nitrite. It is also possible<br />
that in vivo complexes are being formed, such as between heavy metals<br />
ions and the protein, metallothionein, which may be more toxic (or cause<br />
more organ-selective toxicity) than the original, uncomplexed compound<br />
(Wang et al., 1993).<br />
Bioactivation thus mostly concerns the conversion <strong>of</strong> organic chemicals<br />
to more toxic products. On one hand this may result in stable metabolites<br />
that better fit a receptor binding site, resulting in (in principle) reversible<br />
interactions (Mulder, 1992). On the other hand, the metabolites may be<br />
quite reactive, resulting in essentially irreversible effects which are <strong>of</strong><br />
particular concern when they can escape correction, such as neoplasms or<br />
sensitization.<br />
Mechanisms <strong>of</strong> bioactivation<br />
<strong>Industrial</strong> chemicals have widely different structures. Often the preparations<br />
used contain a variable degree <strong>of</strong> impurities, or are mixtures. In this<br />
chapter only the toxicity <strong>of</strong> pure chemicals will be discussed; obviously<br />
when several compounds are present at the same time in a reaction mix or<br />
a commercial product, the final toxicity may be the result <strong>of</strong> complex<br />
interactions between the substituents, which may cause the toxicity to be<br />
more severe (but also much less serious) than expected.
The bioactivation to reactive intermediates by oxidative, cytochrome<br />
P450-mediated metabolism has been extensively studied. So much so, that<br />
it is <strong>of</strong>ten overlooked that conjugation reactions may similarly convert<br />
stable compounds into reactive, electrophilic metabolites (Anders and<br />
Dekant, 1994). This is <strong>of</strong> some practical importance, because many rapid<br />
in vitro toxicity screening tests, e.g. for genotoxicity, include only oxidative<br />
biotransformation capacity (microsomal fractions plus NADPH). In such<br />
screening systems the possibility that, for example, glucuronidation,<br />
sulfation or glutathione conjugation may activate a chemical is not<br />
assessed. Examples <strong>of</strong> bioactivation <strong>of</strong> industrial chemicals by glutathione<br />
conjugation are various halogenated hydrocarbons, while in 2naphthylamine<br />
toxicity glucuronidation may play a role. All in all,<br />
however, little information is available on the role <strong>of</strong> conjugation. As a<br />
consequence, it is unclear at present whether conjugation reactions are <strong>of</strong><br />
major concern for bioactivation <strong>of</strong> industrial chemicals in general. It<br />
certainly seems worth while for reasons more than just scientific curiosity<br />
to include conjugation reactions in test systems. This can be done by using,<br />
for example, intact hepatocytes (or other cells), or by using a mix <strong>of</strong><br />
cosubstrates for conjugation in combination with an S9 fraction (consisting<br />
<strong>of</strong> both cytosol and microsomal fraction). UDP glucuronic acid, a sulfate<br />
activating system, glutathione, acetyl-CoA and S-adenosylmethionine<br />
would cover the major conjugation reactions.<br />
A role <strong>of</strong> bioactivation in the toxicity <strong>of</strong> many chemicals has been<br />
demonstrated. Chemical groups that <strong>of</strong>ten are involved in mutagenic or<br />
carcinogenic effects have been identified (‘alerting groups’). However, as yet<br />
it is still impossible to predict with certainty the carcinogencity <strong>of</strong> a<br />
compound based only on its chemical structure, although a panel <strong>of</strong><br />
experts can make quite good guesses (Wachsman et al., 1993).<br />
In this chapter some <strong>of</strong> the major issues will be illustrated by the<br />
examples vinyl chloride, styrene (versus styrene oxide), benzene,<br />
dichloromethane, chlor<strong>of</strong>orm, 1,2-dibromoethane and 2-naphthylamine.<br />
Vinyl chloride<br />
G.J.MULDER 37<br />
High exposure <strong>of</strong> workers to vinyl chloride in the past has led to the<br />
realization that it may cause neoplasms in man, in particular<br />
haemangiosarcomas in the liver. Vinyl chloride is a genotoxic compound<br />
that acts as initiator <strong>of</strong> various types <strong>of</strong> tumors (Swaen et al., 1987).<br />
The major routes <strong>of</strong> bioactivation <strong>of</strong> vinyl chloride are shown in<br />
Figure 3.1. The most important first step is oxidation by (a) cytochrome<br />
P450 species, resulting in a rather reactive epoxide, which readily<br />
rearranges to chloroacetaldehyde. This may bind to DNA bases, especially<br />
the N6 <strong>of</strong> adenosine or the N4 <strong>of</strong> cytidine, yielding N-ethenoadducts.<br />
Glutathione provides protectionbecause it traps the reactive intermediates
38 METABOLIC ACTIVATION OF INDUSTRIAL CHEMICALS<br />
Figure 3.1 Bioactivation <strong>of</strong> vinylchloride.<br />
formed from vinyl chloride. Furthermetabolism <strong>of</strong> such conjugates leads to<br />
urinary products that can be used tomonitor vinyl chloride exposure in<br />
workers (Guengerich, 1992).<br />
The compound is mutagenic in many in vitro test systems, which require<br />
bioactivation by a microsomal preparation with co-factors for cytochrome<br />
P450. Whether other toxic effects that have been associated with vinyl<br />
chloride exposure in man, such as Raynauds syndrome or acro-osteolysis,<br />
also require bioactivation <strong>of</strong> vinyl chloride is unknown. In addition to its<br />
DNA adduct forming capacity, vinyl chloride also binds covalently to thiol<br />
groups in proteins. It is conceivable that such binding in specific cell types<br />
might lead to non-carcinogenic defects in organ functions.<br />
Styrene and styrene oxide<br />
Styrene metabolism and bioactivation are very similar to that <strong>of</strong> vinyl<br />
chloride: epoxidation by cytochrome P450 is the pathway <strong>of</strong> toxification<br />
(Figure 3.2). It can be detoxified by epoxide hydrolase and glutathione<br />
transferase activity. Mandelic acid excretion in urine can be used for<br />
exposure monitoring in man. Styrene oxide is a direct mutagen in several in<br />
vitro mutagenesis systems and it readily reacts with DNA in vitro.<br />
However, when animals are exposed to styrene in vivo very little if any<br />
DNA binding is observed. Moreover, styrene is not carcinogenic in animal<br />
experiments, although it is a (weak) mutagen in vitro, after bioactivation<br />
(Bond, 1989; Ecetoc, 1992). The explanation most likely is that the styrene
Figure 3.2 Bioactivation <strong>of</strong> styrene.<br />
Figure 3.3 Bioactivation <strong>of</strong> chlor<strong>of</strong>orm.<br />
oxide, generated in vivo inside a cell is such a good substrate for the phase<br />
2 enzymes, epoxide hydrolase and glutathione transferase, that virtually<br />
immediately upon its synthesis, it is further metabolized. Thus, presumably<br />
the build-up <strong>of</strong> an effective concentration in vivo is prevented. Whether<br />
other toxicity <strong>of</strong> styrene in, for example, oesophagus, stomach or<br />
forestomach is related to covalent binding <strong>of</strong> styrene oxide to protein thiol<br />
groups in those tissues is unclear at present.<br />
Styrene is an example <strong>of</strong> a compound <strong>of</strong> which the metabolism<br />
completely goes through a reactive intermediate (the epoxide); yet it does<br />
not cause the cancer that might be expected from its highly mutagenic<br />
metabolite. Accumulation <strong>of</strong> enough <strong>of</strong> this epoxide inside the cells for a<br />
detectable genotoxic effect may require a dose which is acutely toxic, and<br />
therefore can never be tested.<br />
Chlor<strong>of</strong>orm<br />
G.J.MULDER 39<br />
Chlor<strong>of</strong>orm is acutely toxic in the liver and the kidney. This is the result <strong>of</strong><br />
formation <strong>of</strong> a reactive intermediate (Figure 3.3), phosgene, which binds
40 METABOLIC ACTIVATION OF INDUSTRIAL CHEMICALS<br />
avidly to thiol and amine groups in protein. In mice the kidney toxicity is<br />
much more pronounced in males than in females; this sex-difference is due<br />
to the much higher activity <strong>of</strong> the bioactivating cytochrome P450 species in<br />
male mouse kidney than in the females (Pohl et al., 1984). Chlor<strong>of</strong>orm also<br />
increased the tumor incidence in the liver and kidney in some experiments<br />
(Reitz et al., 1990), at dose levels which damaged these organs. However,<br />
there are no indications <strong>of</strong> mutagenicity or genotoxicity in in vitro or<br />
animal in vivo systems. Therefore, most likely the increased tumor<br />
frequency in animals is due to tissue toxicity, leading to increased cell<br />
turnover and a mitogenic stimulus. This is an important distinction, at least<br />
in some countries such as The Netherlands, because for such chemicals a<br />
threshold approach is allowed, whereas for initiating chemicals a linear<br />
extrapolation for carcinogenic risk is used.<br />
Benzene<br />
Benzene presents something <strong>of</strong> a mystery in the evaluation <strong>of</strong> its toxicity<br />
mechanism (Swaen et al., 1989). Exposure to high levels <strong>of</strong> benzene has<br />
been associated with leukaemia in man. However, in vitro it shows little<br />
genotoxicity, and it hardly generates DNA adducts when it is given even at<br />
high dose to animals. A candidate for DNA damage could have been the 1,<br />
4-dihy-droxybenzene (hydroquinone) metabolite, which, however, does not<br />
form DNA adducts readily. Recently a ring-opened metabolite, the<br />
trans,trans-muconic dialdehyde has been proposed as a possible reactive<br />
metabolite <strong>of</strong> benzene (Figure 3.4). Whether it really plays a role in<br />
benzene toxicity is unclear as yet (Kline et al., 1993).<br />
Dichloromethane<br />
Dichloromethane can be metabolized by two pathways, an oxidative and a<br />
conjugative route. Oxidation catalyzed by P450 yields carbon monoxide<br />
(Figure 3.5). The glutathione pathway generates a reactive intermediate,<br />
which is mutagenic and has been implicated in the hepatocarcinogenic<br />
effect <strong>of</strong> dichloromethane in mice. It could be shown that the human liver<br />
has a negligible activity <strong>of</strong> the glutathione transferase involved, so that the<br />
risk for hepatocarcinogenesis in man is virtually non-existent (Green et al.,<br />
1988; Reitz et al., 1989; Dankovic and Bailer, 1994). This example<br />
illustrates how insight into the mechanism <strong>of</strong> bioactivation enables a more<br />
reliable species extrapolation in terms <strong>of</strong> hazard and risk.<br />
1,2-Dibromoethane<br />
This compound can be conjugated with glutathione to form a reactive<br />
thiiranium ion which forms adducts with DNA. This is the reason for the
Figure 3.4 Possible route <strong>of</strong> bioactivation <strong>of</strong> benzene.<br />
Figure 3.5 Bioactivation <strong>of</strong> dichloromethane.<br />
carcinogenic and mutagenic effects <strong>of</strong> 1,2-dibromoethane (Inskeep et al.,<br />
1986).<br />
2-Naphthylamine<br />
G.J.MULDER 41<br />
2-Naphthylamine causes bladder tumors in the dog and man, but not in<br />
mice and rats. The most likely cause is a complicated interplay between<br />
glucuroni dation and urinary pH. In all four species 2-naphthylamine is Nhydroxylated<br />
and subsequently N-glucuronidated. The resulting metabolite
42 METABOLIC ACTIVATION OF INDUSTRIAL CHEMICALS<br />
is excreted in urine. In man and dog the urine is slightly acidic, while in rat<br />
and mouse it is slightly alkaline. Under acidic conditions the glucuronide is<br />
hydrolyzed to generate the hydroxylamine in the bladder. In this case<br />
glucuronidation is not a bioactivation, but rather a targeting<br />
biotransformation: in man and dog the carcinogenic metabolite is targeted<br />
to the bladder, due to the (necessary!) acidic local pH (Kadlubar et al.,<br />
1981).<br />
Conclusions<br />
The above illustrates the importance <strong>of</strong> bioactivation in toxicity <strong>of</strong><br />
industrial chemicals. Is it possible to predict bioactivation from the<br />
structure? As outlined above, in some cases the compound contains<br />
structural elements which make bioactivation to a reactive intermediate<br />
quite likely. Whether it does play a role in toxicity then is still uncertain.<br />
Test systems to detect reactive intermediates depend on, for example, the<br />
availability <strong>of</strong> the radiolabeled compound; in fact, a very high specific<br />
radioactivity is required to detect low levels <strong>of</strong> binding. Alternatively,<br />
radiolabelled glutathione can be used for those reactive intermediates that<br />
readily bind to the thiol group <strong>of</strong> glutathione (Mulder and Le, 1988).<br />
Whether such systems can pick up every relevant toxic reactive<br />
intermediate remains to be seen.<br />
For extrapolation <strong>of</strong> one species to the other it is important to have<br />
insight into the metabolite that is responsible for the toxicity. Therefore, it<br />
is more than just <strong>of</strong> academic interest to know the mechanism <strong>of</strong> toxicity in<br />
safety assessment <strong>of</strong> industrial chemicals. Unfortunately, it is <strong>of</strong>ten not easy<br />
to establish such a mechanism beyond reasonable doubt: it may require too<br />
many rats to feel comfortable about it if we would have to do this for every<br />
chemical used industrially!<br />
References<br />
ANDERS, M.W. (Ed.), 1985, Bioactivation <strong>of</strong> Foreign <strong>Compounds</strong>, Orlando, FL:<br />
Academic Press.<br />
ANDERS, M.W. and DEKANT, W., 1994, Conjugation-dependent Carcinogenicity<br />
and Toxicity <strong>of</strong> Foreign <strong>Compounds</strong>, Orlando, FL: Academic Press.<br />
BOND, J.A., 1989, Review <strong>of</strong> the toxicology <strong>of</strong> styrene, CRC Crit. Rev. Toxicol 19,<br />
227–49.<br />
DANKOVIC, D.A. and BAILER, A.J., 1994, The impact <strong>of</strong> exercise and<br />
intersubject variability on dose estimates for dichloromethane derived from a<br />
physiologically based pharmacokinetic model, Fund. Appl. Toxicol, 22, 20–5.<br />
ECETOC, 1992, Technical report No. 52, Styrene toxicology. Investigations on the<br />
potential for carcinogenicity, Brussels: Ecetoc.
G.J.MULDER 43<br />
GREEN, T., PROVAN, W.M., COLLINGE, D.C. and GUEST, A.E., 1988,<br />
Macro molecular interactions <strong>of</strong> inhaled methylene chloride in rats and mice,<br />
Toxicol. Appl. Pharmacol, 93, 1–10.<br />
GUENGERICH, F.R., 1992, Roles <strong>of</strong> the vinylchloride oxidation products 1chlorooxirane<br />
and 2-chloroacetaldehyde in the in vitro formation <strong>of</strong> etheno<br />
adducts <strong>of</strong> nucleic acid bases, Chem. Res. Toxicol, 5, 2–5.<br />
INSKEEP, P.B., KOGA, N.K., CMARIK, J.L. and GUENGERICH, F.P., 1986,<br />
Covalent binding <strong>of</strong> 1,2-dihaloalkanes to DNA, Cancer Res., 46, 2839–44.<br />
KADLUBAR, F.F., UNRUH, L.E., FLAMMANG, T.J., SPARKS, D., MITCHUM,<br />
R.K. and MULDER, G.J., 1981, Alteration <strong>of</strong> urinary levels <strong>of</strong> the carcinogen,<br />
N-hydroxy-2-naphthylamine, and its N-glucuronide in the rat by control <strong>of</strong><br />
urinary pH, inhibition <strong>of</strong> metabolic sulfation, and changes in biliary excretion,<br />
Chem.-Biol. Interact. 33, 129–47.<br />
KLINE, S.A., ROBERTSON, J.F., GROTZ, V.L., GOLDSTEIN, B.D. and WITZ,<br />
G., 1993, Identification <strong>of</strong> 6-hydroxy-trans,trans-2,4-hexadienoic acid, a novel<br />
ring-opened urinary metabolite <strong>of</strong> benzene, Environm. Hlth Perspect., 101,<br />
310–12.<br />
MULDER, G.J., 1992, Pharmacological effects <strong>of</strong> drug conjugates: is morphine 6glucuronide<br />
an exception? Trends Pharmacol. Sci., 13, 302–4.<br />
MULDER, G.J. and LE, C.T., 1988, A rapid simple in vitro screening test to detect<br />
reactive intermediates <strong>of</strong> xenobiotics. Toxicol. In Vitro, 2, 225–30.<br />
POHL, L.R., GEORGE, J.W. and SATOH, H., 1984, Strain and sex differences in<br />
chlor<strong>of</strong>orm-induced nephrotoxicity. Drug Metab. Disposit., 12, 304–7.<br />
REITZ, R.H., MENDRALA, A.L. and GUENGERICH, F.P., 1989, In vitro<br />
metabo-lism <strong>of</strong> methylene chloride in human and animal tissues, Toxicol.<br />
Appl. Pharmacol, 97, 230–46.<br />
REITZ, R.H., MENDRALA, A.L. and CONOLLY, R.B., 1990, Estimating the risk<br />
<strong>of</strong> liver cancer associated with human exposures to chlor<strong>of</strong>orm using PbPK<br />
modeling, Toxicol. Appl. Pharmacol., 105, 443–59.<br />
SWAEN, G.M.H. et al., 1987, A scientific basis for the risk assessment <strong>of</strong> vinyl<br />
chloride, Regul. Toxicol. Pharmacol, 7, 120–7.<br />
SWAEN, G.M.H. et al., 1989, Carcinogenic risk assessment <strong>of</strong> benzene in outdoor<br />
air, Regul. Toxicol. Pharmacol., 9, 175–85.<br />
WACHSMAN, J.T., BRISTOL, D.W., SPALDING, J., SHELBY, M. and<br />
TENNANT, R.W., 1993, Predicting chemical carcinogenesis in rodents,<br />
Environm. Hlth Perspect., 101, 444–5.<br />
WANG, X.P., CHAN, H.M., GOYER, R.A. and CHERIAN, M.G., 1993,<br />
Nephrotoxicity <strong>of</strong> repeated injections <strong>of</strong> cadmium-metallothionein in rats,<br />
Toxicol. Appl. Pharmacol., 119, 11–16.
4<br />
Sizing up the Problem <strong>of</strong> Exposure Extrapolation:<br />
New Directions in Allometric Scaling<br />
D.BRUCE CAMPBELL<br />
Director International Scientific Affairs, Servier Research and<br />
Development, Slough<br />
Introduction<br />
The evaluation <strong>of</strong> the safety <strong>of</strong> industrial chemicals requires the<br />
administration <strong>of</strong> a range <strong>of</strong> doses to test animals over periods <strong>of</strong> time and<br />
the extrapolation in some meaningful way to man. Various risk assessment<br />
models have been suggested which attempt to measure an uncertainty or<br />
safety factor which can be used to extrapolate to man to obtain an<br />
acceptable daily intake (ADI) (Dourson and Stara, 1983). Other<br />
approaches are also used, such as benchmark dose, the smallest dose which<br />
produces a statistical increase in toxicity over the background level (Crump,<br />
1984), or more frequently the LOEL, the lowest observed dose which<br />
produces an adverse effect, and NOEL, the highest dose at which no<br />
adverse effect is observed. There are difficulties in the interpretation <strong>of</strong><br />
these exposure margins since there is <strong>of</strong>ten little information on: (1) the<br />
slope or intensity <strong>of</strong> the effect, (2) species differences in the sensitivity, (3)<br />
the possibility <strong>of</strong> cumulative or irreversible toxicities, etc. But perhaps the<br />
most important weakness in these estimates is the lack <strong>of</strong> knowledge <strong>of</strong> the<br />
actual circulating levels <strong>of</strong> the chemical(s) in the different species. This<br />
problem is particularly pertinent for industrial chemicals and environmental<br />
pollutants where it may be unethical to administer doses <strong>of</strong> these<br />
compounds to volunteers which are sufficiently high to measure the<br />
kinetics. It is <strong>of</strong> special concern since it is well known that there are large<br />
interspecies differences in the clearance <strong>of</strong> chemicals and that comparison <strong>of</strong><br />
doses in animals, expressed simply in terms <strong>of</strong> mg kg −1 , provides little<br />
information as to the actual exposure likely to occur. This is not surprising<br />
since small animals have relatively faster blood flow and larger organs than<br />
man when expressed as a percentage <strong>of</strong> body weight, and consequently<br />
clearance is more rapid and circulating levels <strong>of</strong> the administered<br />
compound are lower than could be expected during toxicity testing<br />
(Campbell and Ings, 1988).<br />
However since most mammals share similar physiological and<br />
biochemical actions these differences in physiological rates and sizes for
most processes in the mammalian body have been shown to be<br />
proportional to the body weight <strong>of</strong> the animal (Adolph, 1949; Calabrese,<br />
1983; Peters, 1983; Chappell and Mordenti, 1991) and can be related by<br />
allometry, a word from the Greek meaning the measurement (metry) <strong>of</strong><br />
changing size (allo). It has been shown that blood flow, organ size,<br />
metabolic and respiratory rate, and many other physiological and<br />
anatomical variables are related by the general allometric equation<br />
(Boxenbaum, 1982b):<br />
(4.1)<br />
where Y is the function to be measured, W the body weight <strong>of</strong> the animal,<br />
a the coefficient and b the exponent. For mammals, whilst a is different for<br />
each function, b is approximately 0.6–0.8 for rates, flows and clearances, 1.<br />
0 for volumes and organ sizes, and 0.25 for cycles and times. Thus<br />
metabolic rate can be calculated from 7.0·W 0.75 , liver blood flow from<br />
37·W 0.85 , blood weight from 0.055·W 0.99 , and respiratory rate from 0.<br />
019·W 0.26 . Since the blood flows and the weights <strong>of</strong> the liver and kidney,<br />
the two major organs <strong>of</strong> elimination, can be similarly allometrically scaled,<br />
it follows that the same formula could in principle be used for<br />
extrapolation <strong>of</strong> the clearance <strong>of</strong> chemicals between species.<br />
In the past there has been much discussion on the possibility <strong>of</strong><br />
predicting human kinetics and distribution from animal data, using<br />
allometry. For industrial chemicals relatively complex physiological models<br />
have been constructed using this knowledge <strong>of</strong> relative blood flows and<br />
organ size to predict what levels <strong>of</strong> exposure could be expected in man<br />
(Andersen et al., 1984), but little work has been published on comparative<br />
interspecies clearances which will dictate the circulating levels. For drugs,<br />
on the other hand, a number <strong>of</strong> reports have been published on the<br />
rationale for the use <strong>of</strong> allometric scaling <strong>of</strong> kinetics (Dedrick, 1973;<br />
Boxenbaum, 1982b, 1984, 1986; Mordenti, 1985, 1986; Sawada et al.,<br />
1985; Chappell and Mordenti, 1991) but many have been concerned with<br />
its theoretical aspects rather than with its practical use for prediction.<br />
When scaling has been used, the predictions have not always been<br />
accurate, and the method has therefore not had wide usage. This is<br />
unfortunate since the ability to predict what will be the blood levels in man,<br />
without the need to administer the compound, can potentially have many<br />
advantages in drug development and in the safety testing <strong>of</strong> industrial<br />
chemicals where dosing volunteers is <strong>of</strong>ten unacceptable.<br />
Methods<br />
D.BRUCE CAMPBELL 45<br />
A meta-analysis <strong>of</strong> the papers related to this subject has been made from<br />
those published over the last 20 years. Data before this have largely been<br />
rejected due to the poor design <strong>of</strong> the studies or lack <strong>of</strong> analytical
46 SIZING UP THE PROBLEM OF EXPOSURE EXTRAPOLATION<br />
precision. In the main the data have come from drugs but the same general<br />
considerations would hold for environmental chemicals.<br />
Wherever possible the only compounds included in the analysis have<br />
been those where unbound clearance after systemic administration has been<br />
reported, unless it has been shown that there are little interspecies<br />
differences in protein binding or that absorption is known to be complete<br />
in all the animals. In the past these provisos have not always been met,<br />
leading to incorrect interpretation <strong>of</strong> the data. In most reports the<br />
allometric scaling has used results from at least four species but in some<br />
cases up to 11 have been included. Practically this would involve an<br />
enormous resource and would be difficult when many compounds are<br />
being investigated. For this analysis it has been assumed that only one<br />
species will initially be used and the aim <strong>of</strong> this analysis was to find which<br />
single species would provide the best prediction <strong>of</strong> clearance compared to<br />
that found in man.<br />
Three methods have been used using data, wherever possible, from<br />
mouse, rat, rabbit, dog and monkey (macaques) in a total <strong>of</strong> 60<br />
compounds, with human unbound clearances ranging from 4 to 150 909 ml<br />
min −1 .<br />
Simple allometric equation<br />
Figure 4.1 shows a typical allometric relationship for the clearance <strong>of</strong> the<br />
anticancer drug, fotemustine, showing that equation (4.1) can be made<br />
linear for the determination <strong>of</strong> the variables by logarithmically<br />
transforming the body weight (W) and clearances (CL), as shown in<br />
equation (4.2) where the exponent b can be calculated from the slope <strong>of</strong><br />
the linear regression.<br />
(4.2)<br />
From this analysis <strong>of</strong> all the available papers, where this has been<br />
undertaken with more than four species using data taken from 29<br />
compounds, it was possible to show that the mean exponent (b) is<br />
approximately 0.70±0.15 for unbound clearance, but with a range <strong>of</strong> 0.92–<br />
0.28. This mean value is to be expected since it is comparable to the<br />
exponent for the allometric equation relating physiological rates and<br />
clearances to weight as for metabolic rate, body surface area, hepatic and<br />
renal blood flow, etc. Therefore it would seem that even without a specific<br />
knowledge <strong>of</strong> the clearance in a number <strong>of</strong> different species, it could be<br />
assumed that the exponent <strong>of</strong> 0.7 is a common factor for all chemicals, if it<br />
has not been previously determined. The coefficient a can subsequently be<br />
determined for each compound from only one species according to<br />
equation (4.1), and a predictive value for man determined.
Body surface area (BSA)<br />
It has been suggested that the body surface area provides a good measure<br />
<strong>of</strong> overall metabolic rate and that this may be a better measure <strong>of</strong> relative<br />
clearance between species (Chiou and Hsu, 1988). The BSA has therefore<br />
been cal culated for each species using Meehs Formula, BSA=0.103·W 0.67<br />
(Spector, 1956) and the ratio <strong>of</strong> human BSA to animal BSA multiplied by<br />
the animal clearance, to determine the predicted human clearance.<br />
Life span correction<br />
D.BRUCE CAMPBELL 47<br />
Figure 4.1. Allometric scaling <strong>of</strong> Fotemustine clearance compared with the body<br />
weight in various species.<br />
For some drugs, particularly those which are extensively metabolised but<br />
have a low hepatic clearance, such as phenytoin, antipyrine or caffeine<br />
(Boxenbaum, 1982b; Bonati et al., 1984–5), these simple scaling methods<br />
seem to be poorly predictive for man and an allometric correction using<br />
maximum life potential (MLP) has been used to improve the accuracy<br />
(Figure 4.2). Although the allometric approach using body weight alone is
48 SIZING UP THE PROBLEM OF EXPOSURE EXTRAPOLATION<br />
Figure 4.2. Comparison <strong>of</strong> the allometric interspecies scaling for phencyclidine<br />
using: (top) clearance (CL), and (bottom) clearance corrected for maximum life<br />
potential (MLP) in seven species (redrawn from Owens et al., 1987).<br />
valid for many physiological functions it is poorly predictive <strong>of</strong> longevity<br />
or maximum life potential in man. Using a derived equation based on body<br />
weight alone, humans should only live for 26.6 years, clearly an<br />
underestimate. In fact Sacher (1959) has shown that a better measurement<br />
<strong>of</strong> life span can be calculated using not only body weight but also brain<br />
weight (equation (4.3)), and with this correction the MLP for man<br />
becomes 113 years (Boxenbaum and De Souza, 1988).
(4.3)<br />
Simplistically it has been suggested that these differences in longevity can<br />
be explained by the assumption that in any one species there is a<br />
predetermined or fixed amount <strong>of</strong> total ‘body metabolic potential’ and<br />
once this is used up the animal dies (Boddington, 1978). Boxenbaum (1986)<br />
has extrapolated this concept to include intrinsic hepatic metabolism<br />
suggesting that there is a certain quantity <strong>of</strong> ‘hepatic pharmacokinetic<br />
stuff’ per unit <strong>of</strong> body weight available in a life-time which can be<br />
interrelated by the formula:<br />
(4.4)<br />
where CL is the unbound clearance, and c is a constant for each compound.<br />
Thus, the longer the animal lives, the slower this ‘stuff’ is used up.<br />
Examination <strong>of</strong> the data available from 13 disparate compounds<br />
(Table 4.1), where at least four species have been investigated, shows the<br />
MLP correction has produced good results with an exponent b equal to<br />
unity. Thus this would suggest that the relative clearance between species is<br />
directly proportional to their body weight (W) and MLP, and that animal<br />
(CL (A)) and human clearance (CL (H)) can be simply related according to<br />
equation (4.4).<br />
(4.5)<br />
The maximum life potential (MLP) has been calculated for each animal<br />
from Sacher’s formula (equation (4.3)) (mouse=2.7 y, rat=4.7 y, dog=20 y,<br />
rabbit=8 y, monkey=22 y and human=113 y).<br />
For each drug where the appropriate information was available, the<br />
human clearance has been calculated from each species using the above<br />
approaches and compared with that observed (Table 4.2), and the<br />
percentage prediction measured as:<br />
Results<br />
D.BRUCE CAMPBELL 49<br />
The data from 60 different compounds were used in this ongoing analysis<br />
and as could be expected more data were available for the rat (n=47)<br />
compared to mouse (n=27) and dog (n=28), rabbit (n=24), or monkeys<br />
(n=17). In four cases, valproic acid, diazepam, ceftizoxime and<br />
theophylline, different results were found and data have been analysed<br />
separately. For two classes <strong>of</strong> drugs, β-lactams and benzodiazepines, data<br />
from a number <strong>of</strong> compounds were available (n=6 and 12, respectively), but<br />
only mean values were used in this analysis to minimise a class <strong>of</strong><br />
compounds bias in the results.
50 SIZING UP THE PROBLEM OF EXPOSURE EXTRAPOLATION<br />
Table 4.1 Comparison <strong>of</strong> exponential values for b with MLP corrected clearance<br />
(CL u ·MLP=aW b )<br />
From Figure 4.3 it can be seen that for most species the use <strong>of</strong> the simple<br />
exponent 0.7 provided the worst prediction, particularly in the mouse and<br />
dog, which overestimated the human clearance by approximately 600 and<br />
400 per cent, respectively. The rat and rabbit (100–150 per cent) were<br />
better but the monkey was best giving a small overestimate (36 per cent). The<br />
body surface area calculation for most animals gave a better result<br />
particularly for the rat (48 per cent) and monkey (−28 per cent), but the<br />
best method overall is the use <strong>of</strong> the maximum life potential correction<br />
which provided reasonable predictions, within 50 per cent, for all species<br />
with the exception <strong>of</strong> the mouse (89 per cent). The mean accuracy values<br />
only provide part <strong>of</strong> the picture on predictions and the variation, range and<br />
outliers can give additional information on precision and confidence <strong>of</strong> the<br />
analyses. Table 4.3 shows that although there is reasonable accuracy with<br />
the rat, rabbit and dog, the coefficients <strong>of</strong> variations and range <strong>of</strong> values<br />
for these species are large, particularly in the dog, even though the mean<br />
value is reasonable. However for the monkey most estimates <strong>of</strong> human<br />
clearance fall within close proximity to the mean provid ing good<br />
confidence in the data. Similarly the number <strong>of</strong> all compounds which have<br />
a predictability <strong>of</strong> more than 100 per cent error was large for the dog (18 per<br />
cent) and mouse (11 per cent), less for the rat and rabbit, but none were<br />
found for the monkey. In the rat, where the largest number <strong>of</strong> compounds<br />
were examined (n=56), there is a good correlation (r=0.81, p
Figure 4.3. Mean prediction values (percentage error) for human clearance<br />
calculated for various species using: exponent 0.7, body surface area (BSA), and<br />
maximum life potential correction (MLP).<br />
D.BRUCE CAMPBELL 51<br />
For these life span corrections, equation (4.3) has been used to calculate<br />
MLP, but monkeys in captivity, in contract organisations and zoos<br />
(Carmac, 1994), appear to live longer than the calculated 22 years and<br />
ages <strong>of</strong> 35 years are not uncommon. Substituting this longer life span into<br />
the clearance MLP correction improves the mean accuracy to −14 per cent,<br />
but the range increases and 2 per cent <strong>of</strong> compounds now give a prediction<br />
greater than 100 per cent. Attempts to combine predictions from two or<br />
more animals did not improve the accuracy <strong>of</strong> the predictions but did<br />
marginally improve the confidence <strong>of</strong> these values, particularly when the<br />
data from rat and monkey were averaged, from a confidence interval <strong>of</strong><br />
±20 and ±23 for rat and monkey respectively, when used alone, to ±15<br />
when the results were combined.<br />
From this analysis <strong>of</strong> the data it would appear that measurement <strong>of</strong> the<br />
clearance <strong>of</strong> a drug in the monkey together with a correction for MLP<br />
differences, provide the best overall estimate <strong>of</strong> human clearance with the<br />
greatest confidence in the results, although for many compounds the rat or<br />
even the rabbit are good alternatives. The mouse and the dog, on the other<br />
hand, seem to be poorer animal models to extrapolate to human kinetics.<br />
Discussion<br />
There has in the past been a hesitation to use allometric scaling to predict<br />
the clearance in man, but it would appear from this review <strong>of</strong> the literature<br />
that this approach can be used for predictive purposes with an acceptable<br />
degree <strong>of</strong> accuracy, even when the clearance is measured in only one<br />
species. To put this in perspective, if the actual human clearance was 500 ml<br />
min−1 , the predicted clearance using rat or monkey with MLP correction<br />
would be a oximately 300 ml min 1 ppr<br />
− with a 95 per cent confidence,
52 SIZING UP THE PROBLEM OF EXPOSURE EXTRAPOLATION<br />
Table 4.2 Human unbound clearances <strong>of</strong> the compounds used in this analysis
D.BRUCE CAMPBELL 53<br />
a Campbell DB, 1993 unpublished data.<br />
b CL=812 ml min−1
54 SIZING UP THE PROBLEM OF EXPOSURE EXTRAPOLATION<br />
Table 4.3 Interspecies comparisons <strong>of</strong> human clearance predictions expressed as<br />
percentage from observed clearances using maximum life potential corrections<br />
a MLP=22 years.<br />
b MLP=35 years.<br />
c Percentage <strong>of</strong> compounds with a predicted human clearance greater than 100% <strong>of</strong><br />
that observed.<br />
Figure 4.4. Relationship between observed human clearance and that calculated<br />
from the rat using a maximum life potential (MLP) correction (n=56) (—— line <strong>of</strong><br />
identity).<br />
ranging from 240 to 360 ml min− 1.<br />
The monkey appears to be slightly<br />
better than the rat and rabbit in terms <strong>of</strong> the accuracy <strong>of</strong> prediction, and<br />
with a few exceptions may be phylogenetically more acceptable. This is<br />
perhaps not surprising since most studies which have examined species<br />
differences in metabolism indicate that the monkey is more similar to man<br />
compared to the rat (Caldwell, 1981). All the primate data reported, as far<br />
as can be ascertained, have come from the Rhesus or Cynomolgus, Old<br />
World Macaque monkeys. The same considerations may not be true for<br />
New World monkeys, such as the squirrel or marmoset, but few kinetic<br />
comparisons have been made with these species.<br />
In practice, prediction <strong>of</strong> human clearance would involve measuring the<br />
intravenous or intramuscular kinetics, namely the infinite area under the<br />
curve, for each investigatory compound in two to four animals, together<br />
with an estimate <strong>of</strong> the in vitro protein binding in the animal under<br />
investigation and in human plasma, to obtain the free intrinsic clearance
and then multiply the animal clearance by the ratio <strong>of</strong> weight and MLP,<br />
approximately 13 for the rat and 3.5 for a macaque monkey. Of course, as<br />
shown by these data, there can be exceptions, and the monkey and indeed<br />
the rat may not be a suitable species to undertake allometric scaling for all<br />
compounds. However there is an increasing use <strong>of</strong> in vitro systems such as<br />
isolated microsomes, hepatocytes or hepatic slices, to compare the<br />
metabolic pr<strong>of</strong>iles <strong>of</strong> compounds in animals. If undertaken in conjunction<br />
with allometric scaling, pr<strong>of</strong>ound interspecies differences in the rates and<br />
extent <strong>of</strong> metabolism compared to humans could be observed and provide<br />
information on which is the most suitable species to use for scaling. Since<br />
the allometric scaling for volume appears for most compounds to be<br />
directly proportional to body weight with an exponent <strong>of</strong> approximately 1.<br />
0, half-life can also be easily calculated thereby providing all the necessary<br />
kinetic parameters to simulate plasma levels after repeated dosing in man.<br />
With this information the absolute need to undertake kinetic analysis <strong>of</strong><br />
industrial chemicals in volunteers would be reduced since the exposure<br />
calculated by this procedure is considerably better than that employed<br />
presently using uncertainty factors, giving errors in excess <strong>of</strong> 1000 per<br />
cent.<br />
Further studies are <strong>of</strong> course needed to confirm these initial<br />
observations, particularly with those chemicals used in industry or potential<br />
environmental pollutants, but perhaps this re-evaluation shows that<br />
allometry, when correctly used, may well have a practical role in the<br />
evaluation <strong>of</strong> their potential risk to man.<br />
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D.BRUCE CAMPBELL 55<br />
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PART TWO<br />
Reactive industrial chemicals
5<br />
Metabolism <strong>of</strong> Reactive Chemicals<br />
PETER J.van BLADEREN 1,2 and BEN van OMMEN 1<br />
1 TNO <strong>Toxicology</strong>, Zeist<br />
2 Agricultural University, Wageningen<br />
Introduction<br />
For the purpose <strong>of</strong> the present paper, a reactive chemical will be defined as<br />
a strongly electrophilic agent. Such compounds can bind to the numerous<br />
macromolecular targets in the cell, and thus elicit toxic effects. Binding to<br />
DNA can result in mutations or cancer, binding to proteins or membrane<br />
components to cytotoxicity or specific forms <strong>of</strong> toxicity.<br />
A scale could be drawn up for the reactivity <strong>of</strong> electrophiles. However, it<br />
is not certain that those compounds on the high end <strong>of</strong> the scale, i.e. the<br />
most reactive, would also be the most toxic. On the contrary, these<br />
compounds might be expected to react quickly with water and thus not<br />
reach their target molecules. For the purpose <strong>of</strong> classifying the reactivity <strong>of</strong><br />
electrophiles, the most useful is the theory <strong>of</strong> s<strong>of</strong>t and hard acids and bases<br />
(e.g. Commandeur and Vermeulen, 1990). In principle, the preferential<br />
targets for electrophiles can be derived. Furthermore, an electrophile<br />
showing the highest affinity for the relatively hard nitrogen and oxygen<br />
nucleophiles <strong>of</strong> DNA may pose a higher risk for mutations and cancer than<br />
one reacting preferentially with s<strong>of</strong>t sulfur nucleophiles such as found in<br />
proteins and glutathione.<br />
The following classes <strong>of</strong> electrophiles will be discussed: (1) quinones,<br />
which can both arylate as well as cause toxicity through redox cycling; (2)<br />
derivatives with an actual leaving group such as methylene chloride and<br />
ethylene dibromide, and (3) reagents such as isothiocyanates, isocyanates<br />
and α,β-unsatu-rated<br />
ketones and aldehydes.<br />
The enzymes involved in activation and detoxication<br />
To become toxic, almost all <strong>of</strong> the chemicals to which man is exposed,<br />
including the carcinogens, need metabolic activation. The reactive<br />
intermediates that are formed during metabolism are responsible for<br />
binding to cellular macromolecules which very likely elicit the toxic
P.J.VAN BLADEREN AND B.VAN OMMEN 61<br />
response. In general, other biotransformation enzymes can detoxify these<br />
metabolites. Thus, the concentration <strong>of</strong> the ultimate carcinogen, or<br />
toxicant in general, is the result <strong>of</strong> a delicate balance between the rate <strong>of</strong><br />
activation and the rate <strong>of</strong> detoxification. Although toxicological processes<br />
can be much more complex, interindividual differences in susceptibility are<br />
certainly also a result <strong>of</strong> interindividual differences in this balance between<br />
metabolic activation and detoxification.<br />
The enzymes which are to a large extent responsible for the formation <strong>of</strong><br />
reactive metabolites belong to the family <strong>of</strong> cytochromes P-450. However,<br />
for almost all enzymes involved in biotransformation, examples have been<br />
described <strong>of</strong> activation <strong>of</strong> specific classes <strong>of</strong> chemicals. The main classes <strong>of</strong><br />
enzymes involved in detoxifying chemicals which are reactive per se as well<br />
as reactive metabolites are the epoxide hydrolases and the glutathione Stransferases.<br />
NADPH quinone reductase is involved in the reduction <strong>of</strong><br />
quinones.<br />
Epoxide hydrolases<br />
Metabolites which contain an epoxide moiety may undergo hydrolytic<br />
cleavage to less reactive vicinal dihydrodiols. This reaction is catalyzed by<br />
the enzyme epoxide hydrolase (EH), which was first thought to be<br />
exclusively located in the endoplasmic reticulum (microsomal epoxide<br />
hydrolase, mEH; Oesch, 1972). In later studies on the mammalian<br />
metabolism <strong>of</strong> certain alkyl epoxides, the existence <strong>of</strong> a cytosolic EH (cEH)<br />
was demonstrated (Gill et al., 1974). The two forms <strong>of</strong> EH have<br />
complementary substrate specificity, in that many epoxides, e.g. arene<br />
oxides, which are good substrates for mEH are poor substrates for cEH,<br />
and vice versa, e.g. trans-disubstituted oxiranes are good substrates for cEH<br />
but not for mEH (Hammock and Hasagawa, 1983). Other studies have<br />
pointed to the fact that the common nomenclature <strong>of</strong> ‘microsomal’ and<br />
‘cytosolic’ epoxide hydrolase is not semantically precise: metabolic and<br />
immunochemical studies demonstrated the existence <strong>of</strong> membrane-bound<br />
forms <strong>of</strong> cEH (Guenthner and Oesch, 1983), whereas mEH-like activity<br />
was detected in cytosolic fractions <strong>of</strong> human tissue (Schladt et al., 1988).<br />
Glutathione S-transferases<br />
Glutathione is involved in a variety <strong>of</strong> vital cellular reactions. First, a large<br />
number <strong>of</strong> the various classes <strong>of</strong> xenobiotics to which man is exposed—<br />
industrial, therapeutic as well as naturally occurring chemicals—are<br />
metabolized in vivo to reactive intermediates. Such electrophilic<br />
metabolites may bind to cellular macromolecules and thus cause toxicity.<br />
The formation <strong>of</strong> glutathione conjugates, both by spontaneous reaction<br />
between the reactive species and glutathione as well as catalyzed by the
62 METABOLISM OF REACTIVE CHEMICALS<br />
glutathione S-transferases, is the main detoxification mechanism for<br />
electrophiles in mammalian cells (Chasseaud, 1979). Secondly, via<br />
glutathione peroxidase and the glutathione S-transferases, hydrogen<br />
peroxide and organic peroxides are detoxified, yielding glutathione disulfide<br />
as one <strong>of</strong> the products (Prohaska, 1980). Thirdly, glutathione and the<br />
glutathione S-transferases play a role in the biosynthesis <strong>of</strong> such important<br />
endogenous compounds as prostaglandins and leukotriene C4 (Söderstrom<br />
et al., 1985; Ujihara et al., 1988). In fact, in the latter case one may argue<br />
that an endogenous compound is activated by conjugation with<br />
glutathione, since leukotriene C4 is a mediator <strong>of</strong> the adverse reactions<br />
associated with asthmatic attacks (Samuelson, 1988).<br />
The GSTs are a family <strong>of</strong> isoenzymes with broad and overlapping<br />
substrate selectivity. Although membrane-bound forms <strong>of</strong> GST have been<br />
detected (Morgenstern et al., 1988), GST activity is mainly located in the<br />
cytosol. GSTs are dimers <strong>of</strong> subunits and within a dimer, each subunit<br />
functions independently <strong>of</strong> the other (Mannervik and Jensson, 1982). The<br />
GSTs are now known to be a multi-gene family <strong>of</strong> isoenzymes, which can<br />
be divided into four classes (alpha, mu, pi and theta), based on similarity in<br />
structural, physical and catalytic properties <strong>of</strong> their subunits (Ketterer and<br />
Mulder, 1990; Vos and Van Bladeren, 1990). In addition to their crucial role<br />
in catalyzing glutathione conjugation, GSTs may also be important in<br />
intracellular binding and/or transport <strong>of</strong> endogenous and xenobiotic nonsubstrate<br />
ligands (Listowsky et al., 1988).<br />
The glutathione conjugates initially formed from electrophilic species are<br />
further processed via -glutamyltranspeptidase which splits <strong>of</strong>f the<br />
glutamate residue, and dipeptidases which remove the glycine moiety. The<br />
resultant cysteine S-conjugates are then acetylated to give so-called<br />
mercapturic acids which are excreted into the urine (Jakoby, 1980).<br />
Interestingly, mercapturic acids were the first metabolites derived from<br />
xenobiotics to be recognized as such (Baumann and Preusse, 1879).<br />
In recent years it has become increasingly evident that glutathione<br />
conjugation is also involved in the formation <strong>of</strong> toxic metabolites from a<br />
variety <strong>of</strong> chemicals (Monks et al., 1990b). These metabolites display a<br />
wide spectrum <strong>of</strong> toxic effects, ranging from cytotoxicity to genotoxicity.<br />
The various mechanisms elucidated for the toxic action <strong>of</strong> the conjugates<br />
can be grouped as follows: (1) directly toxic glutathione conjugates may be<br />
formed from vicinal and geminal dihaloalkanes, via the formation <strong>of</strong> sulfur<br />
halfmustards; (2) from several types <strong>of</strong> glutathione conjugates active<br />
metabolites may be formed by further metabolic steps: conjugates <strong>of</strong><br />
hydroquinones can be oxidized to give reactive quinones, and conjugates<br />
derived from haloalkenes are transformed into electrophilic species by the<br />
action <strong>of</strong> cysteine conjugate β-lyase. For both hydroquinones and<br />
haloalkenes the selective nephrotoxicity observed is the result <strong>of</strong> the<br />
targeting <strong>of</strong> the conjugates to the kidneys; (3) glutathione conjugates may
P.J.VAN BLADEREN AND B.VAN OMMEN 63<br />
serve as transporting and targeting agents for compounds that react<br />
reversibly with gluathione such as isothiocyanates, isocyanates and α,<br />
βunsaturated<br />
ketones (Van Bladeren, 1988).<br />
Glutathione S-transferase polymorphism<br />
Genetic variation in the expression <strong>of</strong> GST isoenzymes has been studied<br />
almost solely in man. Considerable variation, possibly indicating a<br />
polymorphism, has been observed for the human liver alpha class<br />
isoenzymes. The ratio <strong>of</strong> GSTA1 and GSTA2 subunits, as determined by<br />
HPLC, was found to range from 0.5 to over 10 (Van Ommen et al., 1990).<br />
However, a division into two groups, with average ratios <strong>of</strong> 1.6±0.3 and 3.<br />
8±0.6 could be made, suggesting an alpha class polymorphism. In view <strong>of</strong><br />
the fact that subunits GSTA1 and GSTA2 together make up a major<br />
portion <strong>of</strong> the GST protein in human liver this potential polymorphism<br />
merits further attention. For class mu isoenzymes a clear polymorphism<br />
has been observed in humans: iso-enzyme GSTM1a-1a was found to be<br />
expressed in only 60% <strong>of</strong> the samples analyzed (Board, 1981). In this study<br />
no account was taken <strong>of</strong> the fact that a second mu class isoenzyme,<br />
isoenzyme GSTM1b-1b was also suggested to play a part in this<br />
polymorphism. In a study on the excretion <strong>of</strong> the mercapturate derived<br />
from 1,3-dichloropropene in exposed workers, however, no difference was<br />
observed between mu-positive and mu negative subjects (Vos et al., 1991).<br />
Quinones and their glutathione conjugates<br />
Two modes <strong>of</strong> reactivity can form the basis <strong>of</strong> the toxicity associated with<br />
quinones: (i) their ability to undergo ‘redox cycling’ and to thereby create<br />
an oxidative stress (Kulkarni et al., 1978), and (ii) their electrophilicity<br />
allowing them to react directly with cellular nucleophiles such as protein<br />
and non-protein sulfhydryls (Dierickx, 1983). Since glutathione is the<br />
major non-protein sulfhydryl present in cells, it comes as no surprise that it<br />
is intimately involved in the biological effects <strong>of</strong> quinones. On the one<br />
hand, glutathione can act as a reducing agent, detoxifying quinones by<br />
converting them to hydroquinones with the concomitant formation <strong>of</strong><br />
glutathione disulfide. On the other hand quinone and hydroquinonethioethers<br />
are formed. Recently considerable evidence has been gathered,<br />
indicating that a variety <strong>of</strong> these thioethers possess biological activity<br />
(Dierickx, 1983; Koga et al., 1986).<br />
The target sites for the biological (toxicological) activity <strong>of</strong><br />
quinonethioethers is to a large extent determined by the glutathione moiety:<br />
as will be discussed, the main targets are the kidney (Monks et al., 1985)<br />
and various enzymes using glutathione as a (second) substrate, e.g. the<br />
glutathione S-transferases (Van Ommen et al., 1988). Bromobenzene is
64 METABOLISM OF REACTIVE CHEMICALS<br />
toxic to proximal renal tubules. The nephrotoxic effect <strong>of</strong> o-bromophenol<br />
and bromo hydroquinone was found to be considerably higher, indicating<br />
that these compounds were situated along the main bioactivation route<br />
(Monks et al., 1985). Subsequent elegant work by Monks and Lau has<br />
shown that in fact the nephrotoxicity is caused by the glutathione<br />
derivatives <strong>of</strong> bromohydroquinone (Lau and Monks, 1990). Interestingly,<br />
the relative toxicity <strong>of</strong> the quinoneglutathione conjugates increases as the<br />
extent <strong>of</strong> glutathione addition increases, i.e. the diglutathionyl derivative is<br />
more toxic than the monoconjugate (Monks et al., 1988b). The tissue<br />
selectivity is a consequence <strong>of</strong> their targeting to renal proximal tubule cells<br />
by the brushborder -glutamyl transpeptidase. AT-125, a selective inhibitor<br />
<strong>of</strong> this enzyme in vivo, protects the kidney from the toxic effects <strong>of</strong> the<br />
conjugates. The toxicity <strong>of</strong> these hydroquinone conjugates is apparently<br />
not mediated by cysteine conjugate β-lyase catalyzed formation <strong>of</strong> thiols.<br />
The inhibitor <strong>of</strong> the lyase, amino-oxyacetic acid, had only minor effects on<br />
the extent <strong>of</strong> toxicity, and the putative product, 6-bromo-2,5dihydroxythiophenol,<br />
needed activation by oxidation before it exerted any<br />
biological effect (Monks et al., 1990b). Thus, the effects <strong>of</strong> these<br />
conjugates apparently are a consequence <strong>of</strong> their oxidation to the<br />
corresponding quinones.<br />
Several isomers <strong>of</strong> 2-bromo-glutathionyl as well as the<br />
bromodiglutathionyl hydroquinones were isolated and tested. Instead <strong>of</strong> a<br />
direct correlation <strong>of</strong> toxicity with the electrochemical properties <strong>of</strong> these<br />
compounds, it was found that the diglutathionyl derivative, which is by far<br />
the most toxic, was the most stable to oxidation at pH 7.4 (Monks and<br />
Lau, 1990). The paradox was clarified by Monks and Lau by determining<br />
the oxidation potentials <strong>of</strong> the breakdown products for the mercapturic<br />
acid pathway: hydrolysis <strong>of</strong> the glutathione moiety gives rise to the cysteine<br />
derivative, which is more readily oxidized than the parent compound<br />
(Monks and Lau, 1990). Apparently two detoxication pathways are<br />
possible for these cysteine derivatives: N-acetylation results in formation <strong>of</strong><br />
the mercapturic acid which again is relatively resistant to oxidation, but<br />
oxidative cyclization <strong>of</strong> cysteinylglycine and cysteine derivatives has been<br />
found to give 1,4-benzothiazines, which do not possess any apparent toxic<br />
properties (Monks and Lau, 1990). The action <strong>of</strong> -glutamyl<br />
transpeptidase can thus result in both activation as found for 2bromohydroquinone<br />
derivatives, but also in detoxication as was observed<br />
for 2,5-dichloro-3-(glutathion-S-yl)hydroquinone and 2,5,6-trichloro-3glutathion-S-yl)hydroquinone<br />
(Mertens et al., 1991). The ease with which<br />
the 1,4-benzothiazines are formed is very likely the determining factor in this<br />
case. A similar pathway has been worked out for p-aminophenol, a known<br />
nephrotoxic metabolite <strong>of</strong> acetaminophen (Eckert et al., 1989, 1990).<br />
Bioactivation <strong>of</strong> halogenated benzenes has long been thought to be the<br />
result <strong>of</strong> oxidation to an epoxide. However, recent studies have shown that
P.J.VAN BLADEREN AND B.VAN OMMEN 65<br />
the covalent binding to cellular macromolecules is not only the result <strong>of</strong> the<br />
first oxidative step, but also <strong>of</strong> the second, the formation <strong>of</strong> a quinone or<br />
hydroquinone from the initially formed phenol. The quinone in turn can be<br />
detoxified by glutathione conjugation. However, although glutathione<br />
protects the liver against toxicity due to these quinones, the conjugates are<br />
transported to the kidney and are there activated to new reactive<br />
intermediates. Thus, increasing the relative amount <strong>of</strong> glutathione Stransferases<br />
in this case would not really protect the organism, but merely<br />
change the target organ <strong>of</strong> the active metabolites.<br />
Chemicals with a leaving group<br />
Methylene chloride<br />
Both vicinal and geminal haloalkanes are bioactivated via conjugation with<br />
glutathione. The glutathione-dependent metabolism <strong>of</strong> the important<br />
industrial solvent dichloromethane yields S-chloromethyl-glutathione as the<br />
initial metabolite (Ahmed and Anders, 1976). This intermediate is held<br />
responsible for the carcinogenicity <strong>of</strong> dichloromethane in the mouse.<br />
Interestingly, this compound does not cause tumors in rats, and this has<br />
been related to the fact that the rate <strong>of</strong> metabolism via the glutathione<br />
pathway, catalyzed by the glutathione S-transferases, is much lower in rat<br />
tissue than in mouse tissue. Man has been postulated to resemble the rat in<br />
this respect and is thus presumably safe from the carcinogenic effects <strong>of</strong><br />
methylene chloride (ECETOC, 1988). When it does not react with cellular<br />
macromolecules, the intermediate S-chloromethyl-glutathione is converted<br />
non-enzymatically to S-hydroxymethyl-glutathione, which easily eliminates<br />
formaldehyde and regenerates glutathione (Ahmed and Anders, 1978).<br />
The glutathione S-transferase isoenzyme involved in the formation <strong>of</strong> Schloromethylglutathione<br />
belongs to class theta. Interestingly, a<br />
considerable amount <strong>of</strong> interindividual variation could be observed in a<br />
group <strong>of</strong> 22 individuals (Bogaards et al., 1993).<br />
1,2-Dibromoethane and 1,2-dichloroethane<br />
The vicinal dihaloalkanes are exemplified by 1,2-dibromoethane and 1,2dichloroethane,<br />
which are mutagenic, carcinogenic as well as nephrotoxic<br />
(Van Bladeren et al., 1980; Wong et al. 1982; Guengerich et al., 1984;<br />
Elfarra and Anders, 1985; Cheever et al., 1990). The metabolism <strong>of</strong> these<br />
compounds involves two pathways, cytochrome P-450 dependent<br />
oxidation and glutathione S-transferase catalyzed formation <strong>of</strong> glutathione<br />
conjugates. The oxidative pathway results in chloro- and<br />
bromoacetaldehyde, respectively. These aldehydes are electrophilic and
66 METABOLISM OF REACTIVE CHEMICALS<br />
thought to be responsible for the covalent binding <strong>of</strong> 1,2-dichloro- and 1,2dibromoethane<br />
metabolites to protein (Inskeep and Guengerich, 1984).<br />
Glutathione conjugation results in the formation <strong>of</strong> S-2haloethylglutathione<br />
derivatives, which are sulfur halfmustards and as such<br />
highly reactive metabolites (Jean and Reed, 1989). The formation <strong>of</strong> these<br />
conjugates is catalyzed by the glutathione S-transferases, and both in rat<br />
and man the alpha-class isoenzymes have been found to be the most<br />
efficient in this catalysis (Cmarik et al., 1990).<br />
The glutathione pathway is responsible for the mutagenicity (Van<br />
Bladeren et al., 1980), the DNA-binding (Koga et al., 1986) as well as very<br />
likely the carcinogenicity (Cheever et al., 1990) <strong>of</strong> 1,2-dichloro- and 1,2dibromoethane.<br />
The S-2-haloethylglutathione derivatives are strong<br />
alkylating agents (e.g. Jean and Reed, 1989). Their electrophilicity is<br />
attributable to neighboring-group assistance. The halogen atom is displaced<br />
by the sulfur atom on the next carbon atom, to form a highly reactive<br />
episulfonium ion. The intermediacy <strong>of</strong> this reactive species is supported by<br />
stereochemical studies as well as NMR data (Van Bladeren et al., 1979;<br />
Dohn and Casida, 1987; Peterson et al., 1988).<br />
The relative importance <strong>of</strong> the oxidative and glutathione-dependent<br />
pathway in vivo is difficult to determine, since both pathways give rise to<br />
the formation <strong>of</strong> the same 2-hydroxyethylmercapturate. Using<br />
tetradeutero-1,2-dibromoethane, the ratio <strong>of</strong> the pathways has been<br />
calculated as 4:1 (Van Bladeren et al., 1981b). However, isotope effects<br />
might have a considerable influence on this ratio (White et al., 1983).<br />
The major DNA-adduct derived from 1,2-dibromoethane has been<br />
identified by Guengerich and coworkers to be S-(2-(N7-guanyl)-ethyl)<br />
glutathione (Ozawa and Guengerich, 1983; Koga et al., 1986). In addition,<br />
the structure <strong>of</strong> one <strong>of</strong> several minor adducts was recently found to be S-(2-<br />
(Nl-adenyl)ethyl) glutathione (Dong-Hyun et al., 1990). A series <strong>of</strong> S-2haloethylglutathione<br />
and -cysteine derivatives has been synthesized: all<br />
were found to react with DNA, specifically with guanine residues. As<br />
expected for a mechanism known to involve an intermediate episulfonium<br />
ion, adduct levels were similar for chloro- and bromo-substituted<br />
derivatives. However, in Salmonella typhimurium TA100 a large variation<br />
was observed in the ratio <strong>of</strong> mutations <strong>of</strong> adducts, indicating that the<br />
structure <strong>of</strong> the adduct has a major influence on the mutagenicity<br />
(Humphreys et al., 1990).<br />
Not all vicinal dihaloalkanes seem to give rise to the formation <strong>of</strong><br />
episulfonium ions. Methyl substitution for instance effectively hinders the<br />
mutagenicity through this pathway (Van Bladeren, et al., 1981a) and<br />
studies on 1,2-dibromopropane (Zoetemelk et al., 1986) and hexadeuterol,2-dichloro-propane<br />
(Bartels and Timchalk, 1990) indicate that the<br />
resulting mercapturic acids are only formed through an oxidative pathway.<br />
However, for the heavily used agricultural chemical l,2-dibromo-3-
P.J.VAN BLADEREN AND B.VAN OMMEN 67<br />
chloropropane evidence has been accumulating recently, implicating a<br />
glutathione-mediated activation pathway in the renal and testicular toxicity<br />
associated with this compound (Pearson et al., 1990). Interestingly,<br />
consecutive formation <strong>of</strong> two episulfonium ions can occur, and in fact bis-<br />
DNA-adducts have been identified (Humphreys et al., 1991). 1,2-<br />
Dibromochloropropane could thus cross-link DNA strands as the initial<br />
step leading to cell death.<br />
Isoenzyme selectivity for both primary reactions has been studied<br />
extensively. The alpha and theta class glutathione S-transferases are<br />
responsible for the conjugation <strong>of</strong> EDB both in rats and man. For both <strong>of</strong><br />
these enzymes enormous differences in levels between individuals have been<br />
found, which may be due to genetic differences, but are certainly also<br />
influenced by induction. One might expect individuals with an increased<br />
relative amount <strong>of</strong> glutathione S-transferases to be at increased risk.<br />
Reversible glutathione conjugates acting as transporting<br />
agents<br />
Numerous substrates for glutathione conjugation exist where a formal<br />
addition takes place: both the glutathionyl residue and the hydrogen atom<br />
are added to the acceptor molecule. From a chemical point <strong>of</strong> view, this<br />
reaction should be relatively easily reversible. Of course, the extent <strong>of</strong> the<br />
occurrence <strong>of</strong> the reverse reaction depends on the position <strong>of</strong> the<br />
equilibrium and is influenced by such conditions as the concentration <strong>of</strong><br />
the reactants and the pH. The biological consequences <strong>of</strong> this reaction<br />
sequence would be that the original electrophile is detoxified initially, but<br />
not permanently: it can be released again and thus appear in unexpected<br />
parts <strong>of</strong> the body. The glutathione conjugate serves as a storage or<br />
transport form for the alkylating agent. Systemic effects <strong>of</strong> highly reactive<br />
compounds might be explained in this way.<br />
For both isothiocyanates and isocyanates evidence for this pathway has<br />
been obtained. Benzyl and allyl isothiocyanate are both naturally occurring<br />
compounds that are excreted mainly as mercapturic acids in urine after<br />
administration to rats (Brüsewitz et al., 1977). However, the mercapturate<br />
in urine is unstable under basic conditions and reforms the free<br />
isothiocyanate. The glutathione, cysteine as well as N-acetyl-cysteine<br />
conjugates derived from these isothiocyanates are all toxic in vitro<br />
(Bruggeman et al., 1986, Temmink et al., 1986). In vivo, the fact that the<br />
conjugates are somewhat more unstable in urine probably plays a role in<br />
the effects. Benzyl isothiocyanate is used for the treatment <strong>of</strong> bladder<br />
infections (Brüsewitz et al., 1977), while allyl iso-thiocyanate causes<br />
bladder tumors in male rats (Dunnick et al., 1982).<br />
The extremely reactive and toxic methyl isocyanate, used in the<br />
manufacture <strong>of</strong> carbamate pesticides, was released into the atmosphere in
68 METABOLISM OF REACTIVE CHEMICALS<br />
large amounts during a disaster in 1984. To explain the systemic effects <strong>of</strong><br />
exposure to this compound, Baillie and coworkers hypothesized that these<br />
are mediated by the glutathione conjugates (Pearson et al., 1990). In fact, a<br />
rapid distribution <strong>of</strong> radioactivity throughout the body was found for rats<br />
exposed to 14C-methyl isocyanate vapor (Ferguson et al., 1988), the<br />
glutathione conjugate was identified in bile (Pearson et al, 1990) and the<br />
mercapturic acid was identified as a major urinary metabolite (Slatter et<br />
al., 1991) <strong>of</strong> rats dosed with methyl isocyanate. As was found for the<br />
isothiocyanates, in aqueous solution the synthetic glutathione conjugates<br />
are in equilibrium with the free electrophiles and glutathione: when an<br />
excess <strong>of</strong> cysteine is added to the solution, the corresponding cysteine<br />
conjugate is formed rapidly (Pearson et al., 1990). It should be realized<br />
however, that although thiols are the prime targets <strong>of</strong> iso- thiocyanates and<br />
isocyanates, the reactions with oxygen and nitrogen nucleophiles also<br />
occur and give rise to adducts that are much more stable (Pearson et al.,<br />
1991).<br />
The veterinary drug furazolidone is metabolized to a reactive metabolite<br />
that possesses an α,<br />
β-unsaturated ketone functionality. A reversible, socalled<br />
Michael adduct <strong>of</strong> this metabolite with glutathione was identified<br />
and has been suggested to play a role in the toxic effects <strong>of</strong> furazolidone<br />
(Vroomen et al., 1987). In fact residues <strong>of</strong> this metabolite covalently bound<br />
to microsomal protein could be trapped by an excess <strong>of</strong> mercaptoethanol<br />
and the glutathione conjugate gives rise to covalent binding to microsomal<br />
protein (Vroomen et al., 1988). Similarly, 2-methylfuran is metabolized to<br />
acetyl acrolein. The glutathione conjugate derived from this metabolite is<br />
unstable, and in fact toxicity <strong>of</strong> 2-methylfuran is potentiated by increasing<br />
glutathione levels by the administration <strong>of</strong> the cysteine precursor L-2oxothiazolidine-4-carboxylate<br />
(Ravindranath and Boyd, 1991).<br />
Thus, the reversibility <strong>of</strong> glutathione conjugation reactions warrants<br />
further investigation. The fact that reactive intermediates can be reformed<br />
might have important implications for the explanation <strong>of</strong> effects at sites<br />
distant from the site <strong>of</strong> initial exposure and/or initial conjugation.<br />
Conclusion<br />
Reactive chemicals can be detoxified fairly efficiently by several ubiquitous<br />
biotransformation enzymes. However, numerous cases have been reported<br />
where the initial detoxification is not the end <strong>of</strong> the story. The various<br />
pathways that the initially formed metabolites may undergo can result in<br />
unexpected toxicities at sites distant from the point <strong>of</strong> entry into the body<br />
<strong>of</strong> the electrophilic xenobiotic or the site <strong>of</strong> formation <strong>of</strong> the electrophilic<br />
metabolite.
P.J.VAN BLADEREN AND B.VAN OMMEN 69<br />
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6<br />
Methods for the Determination <strong>of</strong> Reactive<br />
<strong>Compounds</strong><br />
PETER SAGELSDORFF<br />
CIBA-GEIGY Ltd, Basel<br />
Introduction<br />
It has well been recognised for a long time that adverse effects <strong>of</strong> chemicals<br />
are associated with their reactivity whereby many unreactive chemicals are<br />
metabolised in the cell to a reactive intermediate. Reactive intermediates<br />
are generally electrophiles which undergo reactions with cellular<br />
nucleophiles and the toxicological response is <strong>of</strong>ten the consequence <strong>of</strong> the<br />
covalent binding <strong>of</strong> a chemical to cellular macromolecules. This chapter<br />
will provide a brief survey <strong>of</strong> current methods for the determination <strong>of</strong><br />
protein and DNA adducts generated by reactive compounds, and will<br />
discuss some useful applications <strong>of</strong> these technologies.<br />
Source <strong>of</strong> reactive metabolites<br />
Reactive chemicals are generally strong electrophilic agents. These<br />
compounds can be reactive per se (direct electrophiles), such as methylmethanesulphonate,<br />
epoxides or strained lactones. On the other hand,<br />
unreactive chemicals can be enzymatically converted to electrophilic agents<br />
(indirect electrophiles), such as aromatic amines and nitroarenes to the<br />
corresponding nitrenium ions, polycyclic aromatic hydrocarbons to diol<br />
epoxides or N-nitroso compounds to carbenium ions (Figure 6.1; Magee et<br />
al., 1975; Weissburger and Williams, 1975; Lutz, 1979).<br />
Interaction <strong>of</strong> reactive compounds with cellular<br />
constituents<br />
As electrophiles, these reactive compounds undergo reactions with<br />
nucleophiles. The nature <strong>of</strong> the toxicological response is dependent on the<br />
biological macromolecule affected. Reactions with water or glutathione,<br />
two <strong>of</strong> the most abundant cellular nucleophiles, in most cases, lead to an<br />
inactivation <strong>of</strong> the respective reactive compound.
P.SAGELSDORFF 73<br />
Figure 6.1 Examples <strong>of</strong> electrophilic compounds. The arrows indicate the suspected<br />
electrophilic centres (from Lutz, 1979).
74 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS<br />
Figure 6.2 Nucleophilic centres in nucleobases and DNA ‘adduct library’ indicating<br />
the preferential binding sites on guanine for several classes <strong>of</strong> chemicals. The most<br />
reactive targets are indicated by an arrow (from Lutz, 1979; Beach and Gupta,<br />
1992).<br />
The most important nucleophilic centres in proteins are the side chains<br />
<strong>of</strong> the amino acids cysteine, methionine, histidine and tyrosine, and the<br />
amino group <strong>of</strong> the N-terminal amino acid. Reactions <strong>of</strong> electrophiles with<br />
proteins lead to the formation <strong>of</strong> protein adducts. This results in general or<br />
specific cytotoxicity depending on whether the function <strong>of</strong> a particular<br />
protein is disturbed (Lawley, 1976; Brooks, 1977). Adduct formation with<br />
blood proteins can result in the formation <strong>of</strong> immunogens and subsequent<br />
allergenic responses.<br />
Finally, reactions with DNA predominantly occur with the nucleobases<br />
adenine, cytosine, thymine and guanine, whereby the most important<br />
nucleophile in DNA is guanine (Figure 6.2). Adduct formation with<br />
nucleobases in DNA is recognised as a crucial step in the formation <strong>of</strong><br />
mutations and cancer (Lutz, 1979).
Methods for the determination <strong>of</strong> adducts<br />
Adduct formation <strong>of</strong> reactive compounds with protein or DNA can easily<br />
be detected by the use <strong>of</strong> radiolabelled test compounds. However, the<br />
radiolabelled compounds are <strong>of</strong>ten not available. In addition, real exposure<br />
situations and unknown mixtures <strong>of</strong> compounds can not be assessed. For a<br />
number <strong>of</strong> chemicals, therefore, alternative methods for adduct<br />
determinations have been developed during the past.<br />
Reactions with proteins can be assessed by analysing haemoglobin or<br />
albumin adducts. These proteins can easily be isolated in large quantities<br />
(100 mg haemoglobin, 30 mg albumin per ml blood) and with sufficient<br />
purity from the blood <strong>of</strong> treated animals or occupationally exposed<br />
humans. Methods for the determination <strong>of</strong> DNA adducts generally require<br />
a higher sensitivity, since DNA from treated animals or exposed humans is<br />
only available in small amounts (1–2 mg per g tissue, 4 µg from white<br />
blood cells per ml blood).<br />
Protein adducts<br />
Physical methods<br />
Aromatic amines and nitroarenes<br />
The key step in the metabolic activation <strong>of</strong> arylamines to the respective<br />
nitrenium ions involves N-hydroxylation. The N-hydroxylamines can be<br />
further oxidised in erythrocytes to the corresponding nitroso compounds<br />
with a concurrent production <strong>of</strong> methaemoglobin. On the other hand,<br />
nitroarenes can be metabolically reduced to the corresponding<br />
nitrosoarenes. The nitrosoarenes covalently bind to the thiol group <strong>of</strong><br />
cysteine residues and rearrange to give stable sulphinic acid amides.<br />
Mild alkaline treatment can be used to hydrolyse these adducts. The<br />
liberated parent amines can be extracted and analysed by HPLC with<br />
specific detection methods, such as electrochemical or fluorescence<br />
detection. In order to improve the sensitivity <strong>of</strong> the assay, the extracted<br />
adducts can be derivatised with electrophores and analysed by GC with<br />
electron capture detection or by GC/MS (Bailey et al., 1990; Skipper and<br />
Tannenbaum, 1990; Sabbioni, 1992, 1994).<br />
Polycyclic aromatic hydrocarbons<br />
P.SAGELSDORFF 75<br />
Polycyclic aromatic hydrocarbons are oxidised by cytochrome P450 to<br />
epoxides, which are rapidly hydrolysed. However, further oxidation to the
76 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS<br />
Figure 6.3 Modified Edman degradation <strong>of</strong> alkylated N-terminal valine in<br />
haemoglobin (Törnqvist et al., 1986).<br />
respective diol epoxides results in the formation <strong>of</strong> relatively stable<br />
electrophiles which also alkylate cysteine residues in proteins.<br />
Upon mild acid treatment these adducts are liberated as the respective<br />
tetrols. Similar to adducts from aromatic amines, the tetrols can be<br />
extracted and analysed by HPLC with specific detection methods, or by GC<br />
with electron capture detection or by GC/MS after derivatisation with<br />
electrophores (Shugart and Kao, 1985; Weston et al., 1989; Day et al.,<br />
1990).<br />
Alkylating agents<br />
Adducts <strong>of</strong> alkylating agents with the thiol group <strong>of</strong> cysteine, histidine or<br />
the N-terminal amino acids resist alkaline or acid hydrolysis. To determine<br />
the alkylated amino acids the protein is hydrolysed with 6 N HCl and the<br />
amino acids are separated on a anion exchange column. The fractions<br />
containing the alkylated amino acids are derivatised with electrophores and<br />
analysed by GC/MS (van Sittert et al., 1985; Bailey et al., 1987).
Alternatively, the alkylated N-terminal valine <strong>of</strong> haemoglobin can<br />
selectively be cleaved <strong>of</strong>f by a modified Edman degradation with<br />
pentafluorophenyl isothiocyanate (PFPITC). Since alkylation <strong>of</strong> the amino<br />
group favours the reaction, conditions can be selected to exclusively<br />
liberate alkylated N-terminal amino acids whilst leaving the non-adducted<br />
N-terminal valine intact (Törnqvist et al., 1986). The resulting<br />
pentafluorophenyl thiohydantoine (PFPTH) derivative can be extracted and<br />
quantified by GC/MS (Figure 6.3).<br />
Immunological methods<br />
Immunological methods have been developed for the quantification <strong>of</strong><br />
some adducts <strong>of</strong> aromatic amines, polycyclic aromatic hydrocarbons and<br />
alkylating agents. However, these methods involve a couple <strong>of</strong> time<br />
consuming steps for the isolation <strong>of</strong> an appropriate antibody. The<br />
respective haemoglobin adduct has to be chemically synthesised, an animal<br />
has to be immunised with the modified haemoglobin and, later on,<br />
polyclonal antibodies can be isolated from the blood <strong>of</strong> the immunised<br />
animal. In order to produce monoclonal antibodies, which normally have a<br />
better specificity and sensitivity, spleen cells <strong>of</strong> the immunised animal are<br />
fused with myeloma cells and the antibodies can be isolated from the cell<br />
culture.<br />
The methods for the determination <strong>of</strong> adducts include competitive<br />
radioimmunoassays and solid phase assays (ELISA, USERIA). The protein<br />
is partially hydrolysed, adsorbed on a solid surface and treated with the<br />
primary antibody. An anti-antibody which is directed against the primary<br />
antibody, radiolabelled, or conjugated to a fluorescent dye or an indicator<br />
enzyme, is added and the amount <strong>of</strong> bound label is quantified (Santella et al.,<br />
1986; Lee and Santella, 1988).<br />
DNA adducts<br />
Physical methods<br />
Aromatic amines and nitroarenes<br />
P.SAGELSDORFF 77<br />
The hydroxylamines produced by enzymatic hydroxylation <strong>of</strong> aromatic<br />
amines or by reduction <strong>of</strong> nitrosoarenes are further conjugated (Osulphatation,<br />
O-acetylation, O-glucuronidation). The conjugates can<br />
decompose to the respective nitrenium ions which add predominantly to<br />
the C8 <strong>of</strong> guanine.<br />
Similarly to protein adducts <strong>of</strong> these compounds, the adducts can be<br />
liberated from DNA by alkaline hydrolysis or hydrazinolysis, extracted and
78 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS<br />
quantified by HPLC with fluorescence or electrochemical detection. In<br />
order to improve the sensitivity the extracted adducts can be derivatised<br />
with electrophores and analysed by GC/MS (Bakthavachalam et al., 1991;<br />
Lin et al., 1991).<br />
Polycyclic aromatic hydrocarbons<br />
The diol epoxides enzymatically produced from polycyclic aromatic<br />
hydrocarbons mainly adduct at the exocyclic amino group <strong>of</strong> guanine. The<br />
adducts can be liberated from DNA by acid hydrolysis, extracted and<br />
quantified by HPLC with fluorescence or electrochemical detection or by<br />
GC with electron capture detection or GC/MS after suitable derivatisation<br />
with electrophores (Rahn et al., 1982; Shugart and Kao, 1985; Weston et<br />
al., 1989).<br />
Alkylating agents<br />
Alkylating agents mainly alkylate the N7 <strong>of</strong> guanine but also give rise to<br />
the formation <strong>of</strong> other N- and O-alkyl nucleobase adducts. The DNA bases<br />
are liberated by hydrolysis and analysed for the presence <strong>of</strong> adducts by<br />
HPLC with electrochemical detection or they are extracted, derivatised<br />
with electrophores and analysed by GC/MS (Minnetian et al., 1987; Groot<br />
et al., 1994). Some alkylating agents and small epoxides lead to the<br />
formation <strong>of</strong> cyclic nucleobase adducts which exhibit strong fluorescence.<br />
Enzymatic or acid hydrolysis can be used for the liberation <strong>of</strong> the DNA<br />
constituents and the fluorescent adducts can be analysed by HPLC with<br />
fluorescence detection (Fedtke et al., 1990; Steiner et al., 1992a).<br />
Immunological methods<br />
Immunological methods for the determination <strong>of</strong> DNA adducts essentially<br />
follow the procedure as outlined already for protein adducts: generation <strong>of</strong><br />
an antibody, absorption <strong>of</strong> the DNA on a solid surface, incubation with the<br />
antibody and a labelled anti-antibody. However, for the production <strong>of</strong> the<br />
antibody an additional step has to be performed. The immune system<br />
normally does not respond to small molecules. Therefore, the chemically<br />
synthesised base or nucleoside adduct has to be coupled to a carrier protein,<br />
in order to obtain an immunogen (Perera et al., 1986; Santella, 1988;<br />
Poirier, 1993).<br />
Postlabelling<br />
One <strong>of</strong> the most popular assays for determination <strong>of</strong> DNA adducts is the<br />
postlabelling assay. The DNA is enzymatically hydrolysed to the four
natural deoxynucleoside-3′-monophosphates (dNp) and the dNp adducts.<br />
The adducted dNp carrying bulky or aromatic substituents are enriched by<br />
extraction with butanol in the presence <strong>of</strong> a phase transfer agent or by<br />
selective digestion <strong>of</strong> the natural (unadducted) dNp with nuclease P1. The<br />
enriched adducted dNp are labelled with [ 32 P]- or [ 33 P]ATP (Figure 6.4).<br />
Polynucleotide kinase T4 is used to catalyse the transfer <strong>of</strong> the labelled<br />
phosphate group from ATP to the 5′ position <strong>of</strong> the dNp. The labelled<br />
deoxynucleoside-3′,5′-bisphosphates are then separated by multidirectional<br />
TLC on polyethyleneimine coated cellulose. Radioactive impurities and<br />
unused ATP are running to the top with phosphate buffer (D1), whereas<br />
nucleotides carrying aromatic or bulky adducts are retained at or near the<br />
origin. The part containing the impurities and unused ATP is cut <strong>of</strong>f, and<br />
the adducts are chromatographed in D3 (opposite to D1) and D4<br />
(perpendicular to D3) with ammonia and ammonia/ propanol or urea<br />
containing buffers. Adduct spots are visualised and quantified by<br />
autoradiography and Cherenkov counting or by phosphor imaging (Gupta,<br />
1985; Reddy and Randerath, 1986; Beach and Gupta, 1992).<br />
Alternatively, the enriched nucleotide adducts can be chemically<br />
derivatised with fluorescent labels and analysed by HPLC with fluorescence<br />
detection. However, this method does not reach the sensitivity <strong>of</strong> the<br />
radioactive assay (Sharma and Jain, 1991; Jain and Sharma, 1993).<br />
Comparison <strong>of</strong> different methods<br />
P.SAGELSDORFF 79<br />
The methods used for the determination <strong>of</strong> protein and DNA adducts are<br />
summarised in Tables 6.1 and 6.2. Special attention is drawn to the cost <strong>of</strong><br />
equipment and time required for analysis.<br />
HPLC methods with electrochemical or fluorescent detection are<br />
relatively insensitive and only applicable with compounds which are<br />
strongly fluorescent or electrochemically active. Since the costs for the<br />
equipment used and the time consumption are relatively low, these<br />
methods are attractive in certain cases. GC with electron capture detection<br />
or GC/MS <strong>of</strong>fers better sensitivity. However, the method requires<br />
derivatisation. In addition, the costs for the equipment <strong>of</strong> the GC/MS<br />
methods are quite high. Immunoassays are very sensitive, but involve a<br />
number <strong>of</strong> time consuming steps for the preparation <strong>of</strong> an appropriate<br />
antibody, and are only possible if the structure <strong>of</strong> the respective adduct is<br />
known. The postlabelling method for DNA adducts <strong>of</strong>fers the best<br />
sensitivity, with low equipment costs and low to medium time<br />
consumption. However, the standard method only detects bulky or<br />
aromatic adducts.
80 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS<br />
Figure 6.4 Schematic representation <strong>of</strong> the postlabelling assay for determination <strong>of</strong><br />
DNA adducts (Gupta, 1985; Reddy and Randerath, 1986; Beach and Gupta, 1992).<br />
Examples/applications<br />
In the following sections some useful applications <strong>of</strong> adduct<br />
determinations, which have been performed in our laboratory, will be<br />
presented.
Table 6.1 Methods for the determination <strong>of</strong> protein adducts<br />
a An approximate mean sensitivity is given in pmol (=10 −12 mol) adducts/g<br />
haemoglobin.<br />
Table 6.2 Methods for the determination <strong>of</strong> DNA adducts<br />
Lack <strong>of</strong> bioavailability <strong>of</strong> 3,3′-dichlorobenzidine from<br />
diarylide pigments<br />
P.SAGELSDORFF 81<br />
a An approximate mean sensitivity is given in fmol (=10 −15 mol) adducts/mg DNA.<br />
3,3′-Dichlorobenzidine is an important intermediate in the production <strong>of</strong><br />
diarylide pigments and azo dyes. Some <strong>of</strong> these pigments have been tested<br />
in long term studies and shown to exert no specific toxicological effects and<br />
to be not carcinogenic to experimental animals (ETAD Report, 1990).<br />
However, there might be a theoretical hazard after metabolic splitting <strong>of</strong> the<br />
pigments into DCB, a known animal carcinogen (IARC, 1982). DCB and<br />
its N-acetylated metabolite are N-hydroxylated and oxidised to the<br />
corresponding nitroso compound which binds to haemoglobin. Since no<br />
repair <strong>of</strong> haemoglobin adducts occurs, these adducts cumulate during the
82 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS<br />
Figure 6.5 HPLC/ECD pr<strong>of</strong>iles obtained after hydrolysis and extraction <strong>of</strong><br />
haemoglobin samples isolated from an untreated rat (control) and from rats treated<br />
for 4 weeks with DCB (2 mg kg −1 ), Direct Red 46 (160 mg kg −1 ), Pigment Yellow<br />
13 (400 mg kg −1 ) and Pigment Yellow 17 (400 mg kg −l ) as well as <strong>of</strong> commercially<br />
available bovine haemoglobin (Hb-bovine).<br />
life span <strong>of</strong> the erythrocyte. Haemoglobin adduct formation, therefore, was<br />
used to monitor the liberation <strong>of</strong> DCB from diarylide pigments.<br />
Rats were treated by daily oral gavage for 4 weeks with the pigment at<br />
daily dose levels <strong>of</strong> 400 mg kg −1 body weight. As a positive control,<br />
animals were treated accordingly with DCB (2 mg kg −1 ) or with Direct Red<br />
46 (160 mg kg −1 ), asoluble azo dye with known bioavailability <strong>of</strong> DCB.<br />
After termination <strong>of</strong> the treatment, haemoglobin was isolated and<br />
hydrolysed in 0.1 N sodium hydroxide. The liberated DCB and<br />
monoacetyl-DCB were extracted with toluene/2-propanol and analysed by<br />
HPLC with electrochemical detection. With 2 mg DCB kg −1 body weight<br />
DCB and monoacetyl-DCB adducts were clearly detectable, amounting up
to 50 ng g −1 haemoglobin (Figure 6.5). No macromolecular adducts were<br />
detectable in the rats treated with the two diarylide pigments. The limits <strong>of</strong><br />
determination would correspond to a daily DCB dose <strong>of</strong> 0.3–0.5 mg kg −1<br />
body weight, indicating that DCB was not liberated from the pigments at a<br />
determination limit <strong>of</strong> 0.3% <strong>of</strong> the DCB equivalents, whereas the<br />
bioavailability <strong>of</strong> DCB in the rats treated with the azo dye could clearly be<br />
confirmed.<br />
Formation <strong>of</strong> glycidaldehyde from glycidylethers<br />
Bisphenol A diglycidylether (BPADGE) is widely used as component <strong>of</strong><br />
epoxy resins. The chemical reactivity <strong>of</strong> this class <strong>of</strong> compounds is a<br />
prerequisite for their technical use, and accounts for the sensitising,<br />
mutagenic and in some cases carcinogenic properties <strong>of</strong> many epoxy resin<br />
monomers. It was suggested that the metabolic inactivation <strong>of</strong> BPADGE by<br />
hydrolysis <strong>of</strong> epoxides may form an equilibrium with its metabolic<br />
activation by oxidative dealkylation <strong>of</strong> the intact glycidyl side chain<br />
followed by the release <strong>of</strong> glycidaldehyde. Cutaneous treatment <strong>of</strong> mice<br />
with glycidaldehyde led to the formation <strong>of</strong> one major epidermal DNA<br />
adduct which was identified as HMEdA<br />
(hydroxymethylethenodeoxyadenosine, Steiner et al., 1992a). This cyclic<br />
deoxyadenosine adduct is strongly fluorescent and can be quantified by<br />
fluorescence measurements.<br />
In order to investigate the formation <strong>of</strong> glycidaldehyde from BPADGE,<br />
mice were treated with BPADGE (2 mg) and the fluorescent<br />
glycidaldehydeDNA adducts formed in epidermal DNA were compared<br />
with those obtained after treatment with glycidaldehyde (2 mg). After 24–<br />
96 h epidermal DNA was isolated, enzymatically digested to the<br />
deoxynucleoside-3'-monophosphates and analysed for the presence <strong>of</strong><br />
HMEdA by HPLC with fluorescence detection (excitation at 231 nm,<br />
emission at 420 nm). In glycidaldehyde treated mice 166 adducts per 10 6<br />
nucleotides could be detected after an exposure time <strong>of</strong> 24 h (Figure 6.6)<br />
whereas with epidermal DNA from BPADGE treated mice 0.2– 0.8<br />
adducts per 10 6 nucleotides were found. This adduct level would be equal<br />
to a dose <strong>of</strong> 10 µg glycidaldehyde, indicating that, at the most, 1.1% <strong>of</strong> the<br />
glycidaldehyde moiety in BPADGE were bioavailable for DNA-adduct<br />
formation (Steiner et al., 1992b).<br />
Determination <strong>of</strong> reactive compounds in unknown<br />
mixtures<br />
P.SAGELSDORFF 83<br />
A challenging task is the analysis <strong>of</strong> reactive metabolities in unknown<br />
mixtures <strong>of</strong> different compounds. In order to assess the impact <strong>of</strong> chemical<br />
pollution on aquatic organisms, rainbow trouts were continuously exposed
84 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS<br />
to the diluted effluent discharges <strong>of</strong> a chemical production plant for 3<br />
months. The plant produced different dyes and chemicals and the waste<br />
water therefore could be contaminated with a variety <strong>of</strong> aliphatic and<br />
aromatic amines and some cyclic aromatic hydrocarbons. After termination<br />
<strong>of</strong> the treatment, liver and gill DNA from exposed and control trouts was<br />
analysed by [ 32 P]postlabelling for the presence <strong>of</strong> DNA adducts.<br />
The DNA was enzymatically hydrolysed to the nucleotides. Adducted<br />
nucleotides were extracted with butanol in the presence <strong>of</strong> the phase<br />
transfer agent tetrabutylammonium chloride and postradiolabelled with<br />
[ 32 P]ATP and PNK. The labelled nucleotides were separated by<br />
multidirectional TLC with 1.0 M phosphate buffer, pH 6.6, in D1, 0.4 M<br />
ammonia in D3 and 4 N ammonia/propanol (1.2:1) in D4. A final<br />
development in direction D4 with 1.0 M phosphate buffer, pH 6.6, was<br />
used as background clean up.<br />
In the trouts exposed to control water no DNA adducts were detectable,<br />
neither in the livers nor in the gills (Figure 6.7). In contrast, in the trouts<br />
exposed to the highest concentration <strong>of</strong> the waste water, at least 4 DNA<br />
adducts could be found in the livers and in the gills. The overall DNA<br />
adduct level in the exposed trouts was relatively low (1 adduct per 10 8<br />
nucleotides, which indicated only a minimal cancer risk for the exposed<br />
fish.<br />
Limitations<br />
However, the methods presented for adduct determination have their<br />
limitations. For protein adduct determination the most popular method is<br />
by HPLC with electrochemical or fluorescence detection after hydrolysis<br />
and extraction <strong>of</strong> the adducts. This is due to the low cost and time<br />
consumption <strong>of</strong> the method. This method is hampered by the possibility <strong>of</strong><br />
interferences, which can elute in the range <strong>of</strong> the compounds <strong>of</strong> interest. For<br />
an exclusion <strong>of</strong> haemoglobin adducts formation at low levels it is therefore<br />
crucial to obtain additional information about the chromatographic peaks<br />
<strong>of</strong> interest, such as for example, by GC/MS.<br />
DNA adducts are <strong>of</strong>ten assessed by [ 32 P]postlabelling. This method is<br />
limited by low yields <strong>of</strong> the enrichment and labelling procedures and by<br />
choosing the appropriate chromatographic conditions for the resolution <strong>of</strong><br />
the labelled adducts. The lack <strong>of</strong> detectability <strong>of</strong> some DNA adducts,<br />
although they may contain aromatic moieties, enforces the use <strong>of</strong> a positive<br />
standard in order to check for the yield <strong>of</strong> the enrichment and the labelling<br />
reaction, and to check for appropriate chromatographic conditions to<br />
resolve the adducts.
P.SAGELSDORFF 85<br />
Figure 6.6 HPLC/fluorescence analysis <strong>of</strong> epidermal DNA hydrolysates from a<br />
control (a) and a BPADGE treated mouse (b), and UV trace <strong>of</strong> synthetic HMEdAp<br />
and HMEdGp (c).
86 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS<br />
Figure 6.7 TLC chromatograms <strong>of</strong> DNA adducts in gills and livers <strong>of</strong> rainbow<br />
trouts, exposed for 3 months to waste water or control water. Top: control water,<br />
liver DNA (left chromatogram), gill DNA (right chromatogram); bottom: waste<br />
water, liver DNA (left chromatogram), gill DNA (right chromatogram).<br />
Conclusions<br />
Each method, although inherently chemically-specific, has its advantages<br />
and limitations depending on the adduct-type. The continued rapid<br />
development <strong>of</strong> the technologies described for assessing biomarkers should<br />
result in more accurate assessment <strong>of</strong> the intracellular reactions <strong>of</strong><br />
chemicals and thereby provide information about the mechanism <strong>of</strong><br />
toxicity <strong>of</strong> a compound under investigation.
Acknowledgements<br />
Grateful thanks to Drs Markus Joppich and Regula Joppich-Kuhn for<br />
haemoglobin adduct analyses and Dr Sandra Steiner for the development<br />
<strong>of</strong> the fluorescence assay for HMEdAp.<br />
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P.SAGELSDORFF 87<br />
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GREEN, L.C., SKIPPER, P.L., TURESKY, R.J., BRYANT, M.S. and<br />
TANNENBAUM, S.R., 1984, In vivo dosimetry <strong>of</strong> 4-aminobiphenyl in rats via<br />
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GROOT, A.J.L., JANSEN, J.G., VAN WALKENBURG, C.F.M. and ZEELAND,<br />
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GUPTA, R.C., 1985, Enhanced sensitivity <strong>of</strong> 32 P-postlabelling analysis <strong>of</strong> aromatic<br />
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JAIN, R. and SHARMA, M., 1993, Fluorescence postlabelling assay <strong>of</strong> DNA<br />
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LAWLEY, P.D., 1976, Carcinogenesis by alkylating agents, in Searle C.E. (Ed.),<br />
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LEE, M.L. and SANTELLA, M., 1988, Quantitation <strong>of</strong> protein adducts as a<br />
marker <strong>of</strong> genotoxic exposure: immunologic detection <strong>of</strong> benzo[a]pyreneglobin<br />
adducts in mice, Carcinogenesis, 9, 1773–7.<br />
LIN, D.-X., LAY, J.O. JR, BRYANT, M. and KADLUBAR, F.F., 1991, Analysis <strong>of</strong><br />
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MAGEE, P.N., PEGG, A.E. and SWANN, P.F., 1975, Molecular mechanisms <strong>of</strong><br />
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MINNETIAN, O., SAHA, M. and GIESE, R.W., 1987, Oxidation-elimination <strong>of</strong> a<br />
DNA base from its nucleoside to facilitate determination <strong>of</strong> alkyl chemical<br />
damage to DNA by GC/MS with electrophore detection, J.Chromatogr., 410,<br />
453–7.<br />
PERERA, F., SANTELLA, R. and POIRIER, M., 1986, Biomonitoring <strong>of</strong> workers<br />
exposed to carcinogens: immunoassay to benzo[a]pyrene-DNA adducts as a<br />
prototype, J. Occup. Med., 28, 1117–23.<br />
POIRIER, M.C., 1993, Antisera specific for carcinogen-DNA adducts and<br />
carcinogenmodified DNA: applications for detection <strong>of</strong> xenobiotics in<br />
biological samples, Mutat. Res., 288, 31–8.<br />
RAHN, R.O., CHANG, S.S., HOLLAND, J.M. and SHUGART, L.R., 1982, A<br />
fluorimetric HPLC assay for quantitating the binding <strong>of</strong> benzo[a]pyrene<br />
metabolites to DNA, Biochem. Biophys. Res. Commun., 109, 262–8.<br />
REDDY, M.V. and RANDERATH, K., 1986, Nuclease P1-mediated enhancement<br />
<strong>of</strong> sensitivity <strong>of</strong> 32 P-postlabelling test for structurally diverse DNA adducts,<br />
Carcinogenesis, 7, 1543–51.<br />
SABBIONI, G., 1992, Quantitative structure activity relationship <strong>of</strong> the Noxidation<br />
<strong>of</strong> aromatic amines, Chem.-Biol. Interact., 81, 91–117.<br />
SABBIONI, G., 1994, Haemoglobin binding <strong>of</strong> nitroarenes and quantitative<br />
structure activity relationships, Chem. Res. Toxicol, 7, 267–74.<br />
SANTELLA, R.M., 1988, Application <strong>of</strong> new techniques for the detection <strong>of</strong><br />
carcinogen adducts to human population monitoring, Mutat. Res., 205, 271–<br />
82.<br />
SANTELLA, R.M., LIN, C.D. and DHARMARAJA, N., 1986, Monoclonal<br />
antibodies to a benzo[a]pyrene diolepoxide modified protein, Carcinogenesis,<br />
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SHARMA, M. and JAIN, R., 1991, Nuclease P1-mediated fluorescence<br />
postlabelling assay <strong>of</strong> AAF modified DNA model d(TACGTA) and calf-thymus<br />
DNA, Biochem. Biophys. Res. Commun., 177, 151–8.
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SHUGART, L. and KAO, J., 1985, Examination <strong>of</strong> adduct formation in vivo in the<br />
mouse between benzo[a]pyrene and DNA <strong>of</strong> skin and haemoglobin <strong>of</strong> red<br />
blood cells. Environm. Health Perspect., 62, 223–6.<br />
VAN SITTERT, N.J., DE JONG, G., CLARE, M.G., DAVIES, R., DEAN, B.J.,<br />
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haematological effects in workers in an ethylene oxide manufacturing plant,<br />
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SKIPPER, P.L. and TANNENBAUM, S.R., 1990, Protein adducts in molecular<br />
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STEINER, S., CRANE, A.E. and WATSON, W.P., 1992a, Molecular dosimetry <strong>of</strong><br />
DNA adducts in C3H mice treated with glycidaldehyde, Carcinogenesis, 13,<br />
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STEINER, S., HÖNGER, G. and SAGELSDORFF, P., 1992b, Molecular dosimetry<br />
<strong>of</strong> DNA adducts in C3H mice treated with bisphenol A diglycidylether,<br />
Carcinogenesis, 13, 969–72.<br />
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HARRIS, C.C., 1989, Fluorescence and mass spectral evidence for the<br />
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adducts in humans, Carcinogenesis, 10, 251–7.
PART THREE<br />
Pulmonary toxicology <strong>of</strong> industrial<br />
chemicals
7<br />
Studies to Assess the Carcinogenic Potential <strong>of</strong><br />
Man-Made Vitreous Fibers<br />
THOMAS W.HESTERBERG, GERALD R.CHASE,<br />
RICHARD A.VERSEN and ROBERT ANDERSON<br />
Schuller International, Inc., Littleton, CO<br />
Introduction<br />
Man-made vitreous fibers (MMVFs) are a class <strong>of</strong> materials which have<br />
found many applications in both residential and industrial settings. MMVFs<br />
are fibrous inorganic substances that are made primarily from rock, clay,<br />
slag or glass. Sometimes referred to as man-made mineral fibers<br />
(MMMFs), the major classes <strong>of</strong> MMVF are refractory ceramic fibers<br />
(RCFs), fibrous glass, rock (stone) wool and slag wool.<br />
RCF, the smallest category <strong>of</strong> MMVF, represents only about 1–2 per<br />
cent <strong>of</strong> the world production <strong>of</strong> MMVF. It is made by melting Al 2O 3 and<br />
SiO 2 in about equal amounts or by melting kaolin clay and then ‘spinning’<br />
or ‘blowing’ this molten material into fibers. Most RCF is used as a high<br />
temperature furnace insulation. World production <strong>of</strong> RCF in 1990 was<br />
about 80 million 1b. Fibrous glass is the largest category <strong>of</strong> the MMVFs<br />
and is used in insulation, air handling, filtration and sound absorption. The<br />
thermal, acoustical and fire resistant properties <strong>of</strong> these products have led<br />
to their widespread use in a variety <strong>of</strong> residential and commercial<br />
applications. Production <strong>of</strong> fibrous glass in North America in 1989 was<br />
approximately 1.8 million t. Slag and rock wool are composed primarily <strong>of</strong><br />
calcium, magnesium, aluminum and silica. Since 1975, most slag wool has<br />
been produced from the waste slag that resulted from the reduction <strong>of</strong> iron<br />
ore to iron. Rock wool fibers are made from basaltic rocks with additives<br />
such as limestone or dolomite. Slag and rock wool are used in residential<br />
and commercial low and high temperature insulation and in acoustical<br />
ceiling tiles and wall panels. About 75% <strong>of</strong> slag wool production is used in<br />
acoustical ceiling tile manufacture in North America.<br />
Animal studies and epidemiological studies have been conducted to<br />
assess the potential biological effects <strong>of</strong> MMVFs. This research has been<br />
reviewed by the International Agency for Research on Cancer (IARC, 1988),<br />
the International Programme on Chemical Safety (IPCS, 1988), and the US<br />
Environmental Protection Agency (Vu, 1988). These reviews are consistent
92 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
in the judgement that chronic inhalation studies <strong>of</strong> airborne fibers provide<br />
the best model for assessing potential risk to man (McClellan et al., 1992).<br />
In assessing the carcinogenic risks <strong>of</strong> exposure to any possible<br />
occupational hazard, research is pursued through several different scientific<br />
techniques. Studies <strong>of</strong> mortality (analysis <strong>of</strong> death rates) are used to<br />
evaluate the potential carcinogenicity associated with direct human<br />
exposure. Animal exposure studies are used to not only evaluate the<br />
potential carcinogenicity but to also investigate the mechanisms <strong>of</strong> disease<br />
development. <strong>Industrial</strong> hygiene and engineering studies are used for<br />
quantifying exposures.<br />
Epidemiological studies<br />
By general agreement among experts (IARC, 1988; IPCS, 1988), two major<br />
historical cohort studies are considered to have comprehensively addressed<br />
the mortality experience <strong>of</strong> workers engaged in the production <strong>of</strong> FG, rock<br />
wool and slag wool: a European study conducted by the International<br />
Agency for Research on Cancer (IARC), and a University <strong>of</strong> Pittsburgh<br />
study conducted in the USA. The discussion here will concentrate on those<br />
two studies. For a summary <strong>of</strong> other studies, the reader is referred to the<br />
IARC review (IARC, 1988). There are no published reports <strong>of</strong> the<br />
mortality experience <strong>of</strong> RCF workers. Epidemiological studies <strong>of</strong> workers<br />
engaged in the manufacture <strong>of</strong> all major classes <strong>of</strong> MMVF are underway or<br />
are continuing. Morbidity studies <strong>of</strong> the respiratory health <strong>of</strong> workers are<br />
not discussed here.<br />
The IARC study<br />
IARC researchers reported their study at the WHO Occupational Health<br />
Conference on the Biological Effects <strong>of</strong> Man-Made Mineral Fibres at<br />
Copenhagen in 1982, with a follow-up in 1986 (Simonato et al., 1987). The<br />
updated study is also published in the Scandinavian Journal <strong>of</strong> Work,<br />
Environment & Health, Volume 12, Supplement 1, 1986. The mortality <strong>of</strong><br />
23609 workers (2836 deaths) employed in 13 European factories engaged<br />
in the production <strong>of</strong> MMVF (including 11 852 fibre glass production<br />
workers at six plants in five countries and 10 115 rock wool/slag wool<br />
production workers at seven plants in four countries) has been studied<br />
(Saracci et al., 1984) and updated (Simonato et al., 1987). The authors<br />
reported an ‘excess <strong>of</strong> lung cancer among rock-wool/slag workers<br />
employed during an early technological phase before the introduction <strong>of</strong><br />
dust-suppressing agents’, and concluded that ‘fiber exposure, either alone or<br />
in combination with other exposures, may have contributed to the elevated<br />
risk’. The authors also reported that ‘no excess <strong>of</strong> the same magnitude was<br />
evident for glass-wool production, and the follow-up <strong>of</strong> the continuous-
filament cohort was too short to allow for an evaluation <strong>of</strong> possible longterm<br />
effects’. It was also noted that ‘there was no evidence <strong>of</strong> an increased<br />
risk for pleural tumors or non-malignant respiratory diseases’. An update <strong>of</strong><br />
this study is underway.<br />
The University <strong>of</strong> Pittsburgh study<br />
This study was also reported at the WHO Occupational Conference on<br />
Biological Effects <strong>of</strong> Man-Made Mineral Fibres at Copenhagen in 1982 and<br />
the follow-up Conference in 1986 (Simonato et al., 1987). Subsequent to<br />
the 1986 Conference, additional analyses were completed and included in<br />
the manuscript published for the proceedings (Enterline et al., 1987). The<br />
study has been updated and published (Marsh et al., 1990). The University<br />
<strong>of</strong> Pittsburgh researchers’ comprehensive mortality review <strong>of</strong> more than<br />
16000 workers— many with long-term exposure up to 40 years—was<br />
undertaken at 17 US fiber-glass, rock-wool and slag-wool manufacturing<br />
plants, including 14800 fiber-glass workers in 11 plants. The original<br />
report, given in 1982, covered the mortality experience from the 1940s to<br />
the end <strong>of</strong> 1977. The same group <strong>of</strong> workers was followed through 1982<br />
(reported in October 1986, with additional analyses available in June<br />
1987). The June 1987 report contained, for the first time, local area<br />
mortality statistics for each <strong>of</strong> the plants as the basis for studying the<br />
mortality experience. Experts agree that, barring unusual circumstances,<br />
local area comparisons are most appropriate. The study has been further<br />
updated through 1985, with publication in 1990. For respiratory cancer, in<br />
the latest update there was a small but statistically significant increase for<br />
fiber glass production workers. However, aside from the issue <strong>of</strong><br />
uncontrolled potential confounding, the study provides no evidence to date<br />
that respiratory cancer mortality is related to fiber glass exposure. There<br />
was a somewhat larger statistically significant excess <strong>of</strong> respiratory cancer<br />
mortality reported for slag wool and rock wool production workers. The<br />
absence <strong>of</strong> any clear exposure-response relationship for any <strong>of</strong> the fiber<br />
groups studied led the authors to conclude that ‘overall, the evidence <strong>of</strong> a<br />
relationship between exposure to man-made mineral fibers and respiratory<br />
cancer appears to be somewhat weaker than in the previous update’.<br />
Consistent with the IARC study, no increase in the occurrence <strong>of</strong><br />
mesothelioma has been observed in this cohort. This study has now been<br />
expanded to include well over 30000 workers from 14 fiber-glass and six<br />
rock wool and slag wool facilities.<br />
Other epidemiological studies<br />
T.W.HESTERBERG ET AL. 93<br />
In addition to the two major studies highlighted above, a number <strong>of</strong> other<br />
studies have been conducted as well. Many <strong>of</strong> them widely overlap these
94 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
major studies, comprise sub-groups within them, or represent smaller<br />
worker populations outside <strong>of</strong> them.<br />
A Canadian study was reported by Shannon at the WHO Occupational<br />
Health Conference on Biological Effects <strong>of</strong> Man-Made Mineral Fibers at<br />
Copenhagen in 1982 and 1986 (Shannon et al., 1987). It followed 2557<br />
male workers at a Canadian glass wool plant through 1977 and was later<br />
updated to extend the follow-up to the end <strong>of</strong> 1984. In the updated study,<br />
the authors reported a statistically significant excess <strong>of</strong> lung cancer. In<br />
discussing this excess, the authors concluded that the interpretation <strong>of</strong> the<br />
information was difficult since there was no relationship between the<br />
excess <strong>of</strong> lung cancer and the length <strong>of</strong> time since first exposure to the<br />
fibrous glass manufacturing environment.<br />
Two recent case-control studies have addressed the lung cancer mortality<br />
<strong>of</strong> FG and slag wool production workers. Chiazze et al. (1992) have<br />
investigated the potential impact <strong>of</strong> confounding factors such as smoking<br />
and other occupational exposures for workers at the oldest and largest US<br />
fiber glass manufacturing facility. In particular, Chiazze helped clarify the<br />
heavy smoking patterns in those workers and verified the large impact that<br />
smoking has on their lung cancer experience. Wong et al., (1991)<br />
investigated the potential impact <strong>of</strong> smoking on the lung cancer deaths at<br />
nine US slag wool manufacturing plants. Wong also found heavy smoking<br />
among the slag wool workers and advanced the understanding <strong>of</strong> the<br />
modest increase in lung cancer seen in the historical cohort studies cited<br />
above.<br />
Users <strong>of</strong> MMVFs generally have experienced mixed exposures, making<br />
the study <strong>of</strong> any potential health effects <strong>of</strong> MMVF difficult, if possible at<br />
all. For example, in a study <strong>of</strong> Swedish construction workers, Engholm et al.<br />
(1987) discussed the difficulty caused by overlapping <strong>of</strong> reported exposures<br />
to asbestos and MMVFs. In addition, essential employment and exposure<br />
histories for users <strong>of</strong> MMVFs are lacking.<br />
The mortality studies <strong>of</strong> FG workers, while showing a small but<br />
statistically significant increase in lung cancer, have failed to show any<br />
consistent relationship with exposure to FG (i.e. no dose-response<br />
relationships have been found). It is recognized that uncontrolled<br />
occupational and/or non-occupational confounding factors may be<br />
associated with the slight increase. The IARC review (IARC, 1988)<br />
concluded that there is ‘inadequate evidence’ for carcinogenicity in<br />
humans. Other reviews have reached similar conclusions. In addition,<br />
reports subsequent to the IARC review have further clarified potential<br />
confounding factors and, if anything, shown weaker evidence <strong>of</strong> a<br />
relationship between exposure and lung cancer.<br />
The cohort mortality studies <strong>of</strong> rock wool and slag wool workers have<br />
shown a somewhat larger statistically significant excess <strong>of</strong> lung cancer<br />
deaths, but have also provided no clear dose-response relationship with
fiber exposure. While the IARC review (IARC, 1988) concluded that there<br />
is ‘limited evidence’ for carcinogenicity in humans, reports subsequent to<br />
the IARC review have further clarified potential confounding factors and,<br />
if anything, shown weaker evidence <strong>of</strong> a relationship between exposure and<br />
lung cancer.<br />
Experimental studies<br />
Toxicologic studies <strong>of</strong> MMVFs have been conducted in both in vitro and in<br />
vivo systems. In addition, the physical and chemical characteristics thought<br />
to correlate with toxicity have been examined. The in vitro studies have<br />
been conducted using cells from the lungs <strong>of</strong> animals as well as bacterial<br />
and cell lines. Two categories <strong>of</strong> whole-animal studies have been reported:<br />
studies using artificial methods to implant high concentrations <strong>of</strong> fibers in<br />
the abdomen, pleura or trachea <strong>of</strong> animals; and inhalation studies <strong>of</strong><br />
maximum tolerated doses and multiple dose levels <strong>of</strong> fibers.<br />
Cell culture studies<br />
T.W.HESTERBERG ET AL. 95<br />
The use <strong>of</strong> cell culture systems for studying the toxic effects <strong>of</strong> fibers has<br />
been recently reviewed (Hesterberg et al., 1993a). A number <strong>of</strong> studies<br />
have shown that fiber length and diameter are important in determining<br />
the toxicity <strong>of</strong> mineral fibers <strong>of</strong> various chemical compositions to cells<br />
grown in culture (Chamberlain et al., 1979, Tilkes and Beck, 1980;<br />
Hesterberg and Barrett, 1984; Hesterberg et al., 1993a; Hart et al., 1994).<br />
Chemical composition has also been shown to be critical to the toxicity <strong>of</strong><br />
fibers to rat tracheal epithelial cells (Ririe et al., 1985) and human<br />
bronchial epithelial cells grown in culture (Kodama et al., 1993). MMVFs<br />
have also been shown to induce neoplastic transformation (Hesterberg and<br />
Barrett, 1984; Poole et al., 1986) and genetic damage to cells in culture<br />
(Sincock and Seabright, 1975; Oshimura et al., 1984). Cell culture models<br />
are important for understanding the mechanisms <strong>of</strong> fiber toxicity and, with<br />
further development, have potential for use as part <strong>of</strong> a battery <strong>of</strong> shortterm<br />
screening tests to assess the toxic and tumorigenic potential <strong>of</strong><br />
mineral fibers. However, it was recently shown that cytotoxicity <strong>of</strong> different<br />
compositions MMVFs to Chinese hamster ovary (CHO) cells in culture did<br />
not correlate with the in vivo toxicity <strong>of</strong> theses MMVFs (Hart et al., 1994).<br />
This may be related to CHO cells being aneuploid, preneoplastic and not a<br />
normal target cell for fiber toxicity in vivo. Future in vitro studies <strong>of</strong> MMVF<br />
toxicity should focus on the use <strong>of</strong> cell types that represent the relevant<br />
target tissues, and cells should be as close to normal as possible.
96 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
Implantation studies<br />
Using various types and dimensions <strong>of</strong> fibers, researchers have studied the<br />
effects <strong>of</strong> ‘artificially’ exposed animals by surgically implanting fibrous<br />
material in the pleural (chest) and abdominal cavities <strong>of</strong> laboratory<br />
animals, and by injecting fibers directly into the trachea (Pott et al., 1987;<br />
Stanton et al., 1981). Those studies have shown that high levels <strong>of</strong> most<br />
fibrous materials <strong>of</strong> certain dimensions, regardless <strong>of</strong> their physical or<br />
chemical makeup, can induce tumors in laboratory animals. From these<br />
study results, scientists have also hypothesized that biological activity<br />
correlates with fiber length and diameter, since ‘long, thin’ fibers are the<br />
most active. The actual chemical composition appears to play only a minor<br />
role, if any, in such ‘artificial exposure’ experiments (Stanton et al., 1981).<br />
Injection <strong>of</strong> fibers bypasses the normal defense mechanisms <strong>of</strong> the lung<br />
and can produce abnormal fiber distribution, fiber clumping, and overload<br />
doses (McClellan et al., 1992). Furthermore, when fibers are injected into<br />
the pleura or peritoneum <strong>of</strong> an animal, leaching, degradation,<br />
fragmentation or any other transformations are unlikely to be the same as<br />
after inhalation. The weaknesses <strong>of</strong> intracavitary injection studies <strong>of</strong><br />
fibrous materials limit their relevance for human risk assessment (IPCS,<br />
1988; Vu, 1988; Dement et al., 1990; WHO, 1992; McClellan et al.,<br />
1992).<br />
Recent animal inhalation studies <strong>of</strong> MMVFs<br />
Inhalation is the only natural route <strong>of</strong> exposure for fiber entry and<br />
distribution to the target organs in man. Animal inhalation studies are<br />
more relevant than intracavity administration studies for risk assessment<br />
because the exposure conditions <strong>of</strong> inhalation experiments more closely<br />
approach the circumstances <strong>of</strong> human exposure.<br />
Background<br />
In June 1988, a series <strong>of</strong> inhalation studies was initiated at Research and<br />
Consulting Company (RCC) in Geneva, Switzerland, to evaluate the<br />
biological effects <strong>of</strong> different compositions <strong>of</strong> MMVF. These included<br />
RCFs, common insulation fiber glass, and rock and slag wool fibers. These<br />
studies used recently perfected state-<strong>of</strong>-the-art technologies for fiber sizeseparation,<br />
fiber l<strong>of</strong>ting and nose-only inhalation exposure. A more<br />
detailed description <strong>of</strong> the techniques used and the results from these<br />
studies are found elsewhere (Hesterberg et al., 1991, 1993b; Mast et al.,<br />
1995 McConnell et al., 1994). The animal models selected were those with<br />
demonstrated capacity to develop asbestosrelated disease following<br />
inhalation exposure. The studies were conducted in accordance with
standard techniques for chronic toxicity/carcinogenicity studies, including<br />
dose and latency considerations. Rats were exposed for 2 years and<br />
hamsters for 18 months. The animals were observed for their lifetime or until<br />
20% survival <strong>of</strong> the test group was reached. Positive and shamexposed<br />
negative controls were included in the protocol.<br />
Fiber aerosols<br />
In designing these animal inhalation studies, the techniques <strong>of</strong> fiber<br />
preparation, aerosolization, exposures, measurement, quantification and<br />
determination <strong>of</strong> actual target organ dose were critical factors. Fiber<br />
dimensions that permitted deposition into the distal lung regions (i.e.<br />
respirable fibers) for the model used were selected. The characteristics <strong>of</strong> the<br />
fiber aerosol in actual work areas for man was an important consideration<br />
in determining experimental exposure. For example, an average fiber size<br />
<strong>of</strong> 1×20 µm has been measured during simulated RCF work practices. The<br />
critical need to use fibers pre-selected for their size and to verify the actual<br />
size distributions <strong>of</strong> the fiber exposure aerosol was met throughout the<br />
study. Non-fibrous particles (shot) in the aerosol were reduced to the<br />
maximum extent possible. Furthermore, fiber preparation, handling and<br />
aerosolization did not alter the physical-chemical characteristics <strong>of</strong> the<br />
fiber, since as will be discussed later, these are known to be critical<br />
determinants <strong>of</strong> fiber toxicity.<br />
Nose-only rather than whole-body exposure was used for several<br />
reasons, including the impossibility <strong>of</strong> preparing the huge quantities <strong>of</strong><br />
specially sized fibers that would be required for 2 years <strong>of</strong> whole-body<br />
exposure. Additionally, nose-only exposure levels permitted better control<br />
<strong>of</strong> exposure levels and host entry.<br />
Selection <strong>of</strong> exposure concentrations<br />
It was important that at least three exposure concentrations be used in the<br />
chronic inhalation study in order to assess the dose-response relationships<br />
<strong>of</strong> any induced changes. The highest concentration selected was the<br />
‘Maximum tolerated dose’ (MTD), while lower concentrations were 50 per<br />
cent <strong>of</strong> the MTD and multiples <strong>of</strong> the projected occupational and<br />
environmental exposure levels.<br />
Experimental design, time lines<br />
T.W.HESTERBERG ET AL. 97<br />
Groups <strong>of</strong> three or six randomly selected animals from each exposure<br />
group were killed at 3, 6, 12, 18 and 24 (rats only) months to follow the<br />
progression <strong>of</strong> histopathological changes and to determine lung fiber<br />
burdens. An additional six ‘recovery’ animals were removed from each
98 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
exposure group at 3, 6, 12 and 18 (rats only) months and held without<br />
further treatment until the end <strong>of</strong> the exposure period, when they were<br />
killed to assess progression or regression <strong>of</strong> lung lesions and lung retention<br />
and clearance <strong>of</strong> fibers after cessation <strong>of</strong> exposure. To assure quality<br />
control, the l<strong>of</strong>ting technique and exposure level were consistently<br />
monitored during the study by both gravimetric measurement and fiber<br />
counting techniques. The terminal sacrifice was carried out when only 20 per<br />
cent <strong>of</strong> the animals survived. A complete necropsy was performed on each<br />
animal. Gross pathological examination and diagnoses were performed<br />
using a dissecting microscope. Uniform sections <strong>of</strong> the left lung and right<br />
diaphragmatic lobe were embedded in paraffin, cut at a thickness <strong>of</strong> 4 mm<br />
and replicate sections were routinely stained with hematoxylin and eosin<br />
(H&E) and Masson-Goldner’s trichrome stain for collagen staining to<br />
assess the presence <strong>of</strong> lung fibrosis. In addition, sections were made from<br />
all grossly visible lesions from that and other portions <strong>of</strong> the lung.<br />
Proliferative lesions <strong>of</strong> the pulmonary parenchyma were designated as<br />
bronchoalveolar hyperplasia, pulmonary adenoma or adenocarcinoma.<br />
Other types <strong>of</strong> lesions, including those in the pleura were noted where<br />
appropriate. All research and analyses were conducted using good<br />
laboratory practices.<br />
Lung fiber burden<br />
Immediately after necropsy, the infracardiac lobe <strong>of</strong> each animal’s lung was<br />
removed and frozen for later analysis <strong>of</strong> lung fiber burden. To recover<br />
fibers from the lung, the tissue was rapidly dehydrated with acetone and<br />
ashed using a low-temperature process. Recovered fibers were dispersed in<br />
distilled water and examined using scanning electron microscopy. Number,<br />
dimensions and other physical characteristics <strong>of</strong> the inhaled lung fibers<br />
were determined, and reported as fibers per mg <strong>of</strong> dry lung weight.<br />
Results from recent animal inhalation studies <strong>of</strong> MMVFs<br />
Refractory ceramic fibers<br />
In the first RCC studies, rodents were exposed to the MTD <strong>of</strong> the sizeselected<br />
RCF test fiber, 30 mg m −3 and approximately 200–250 fibers cm<br />
−3 . Rats were exposed for 6 h per day, 5 days a week to aerosols containing<br />
one <strong>of</strong> four different types <strong>of</strong> RCF: kaolin, RCF 1; zirconia, RCF 2; high<br />
purity kaolin, RCF 3; and ‘after service’ (a kaolin based ceramic fiber<br />
containing 27% crystalline silica that had previously been exposed to high<br />
temperature), RCF 4. Hamsters were exposed to only kaolin RCF fibers.<br />
Positive controls (chrysotile asbestos) and negative controls (filtered air)
Table 7.1 Summary <strong>of</strong> lung pathology findings in RCF hamster inhalation study<br />
a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter
100 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
Table 7.2 Summary <strong>of</strong> lung pathology findings in RCF MTD (30 mg m−3) rat<br />
inhalation<br />
a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter
asbestos (10 mg m −3 ) were included in the study. The crocidolite exposure<br />
had to be stopped at 10 months due to excessive mortality resulting from<br />
lung toxicity. The results are summarized in Table 7.5. Crocidolite<br />
exposure resulted in lung fibrosis, a significant increase in lung tumors, and<br />
a single mesothelioma. Rock wool, but not slag wool, exposure at 16 and<br />
30 mg m −3 resulted in minimal lung fibrosis. However, neither rock wool<br />
nor slag wool exposure resulted in mesotheliomas or a significant increase<br />
in lung tumors.<br />
Lung burden analyses<br />
T.W.HESTERBERG ET AL. 101<br />
Table 7.3 Combined summary <strong>of</strong> lung pathology findings: chrysotile and RCF1<br />
from RCF MTD study in rats; and RCF multidose study in rats<br />
a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter
102 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
Table 7.4 Summary <strong>of</strong> lung pathology findings in fibrous glass inhalation study in<br />
rats<br />
a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter 5 µm, and a diameter 10 µm in length in the lung were similar for each <strong>of</strong> the<br />
different MMVF types (Figures 7.2(a) and 7.2(b)). However, greater<br />
numbers <strong>of</strong> long fibers (>20 µm. long) were found in the lungs <strong>of</strong> rats<br />
exposed to RCF 1 and MMVF 21 (rock wool) than for the other fiber<br />
types (Figure 7.2(c)). Even though lung levels <strong>of</strong> long MMVF 21 fibers<br />
were higher than long RCF 1 fibers, lung fibrosis occurred much later for<br />
MMVF 21 (18 vs 6 months for RCF 1) and no mesotheliomas or significant<br />
increase in lung tumors were observed for MMVF 21. This indicates that<br />
the lung pathogenic potential <strong>of</strong> a fiber may be determined by more than<br />
dose and dimension.
T.W.HESTERBERG ET AL. 103<br />
Table 7.5 Summary <strong>of</strong> lung pathology findings in rock and slag wool inhalation<br />
study in rats<br />
a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter
104 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
disappearance <strong>of</strong> long fibers. More studies are required to determine if a<br />
fiber’s ability to be leached is a critical determinant <strong>of</strong> its ultimate toxicity<br />
to the lung.<br />
Figure 7.1 Length distributions (a) and diameter distributions (b) <strong>of</strong> fibers from the<br />
lungs <strong>of</strong> rats exposed for 13 weeks to the five different MMVFs in the chronic<br />
inhalation studies. To permit clearance <strong>of</strong> the upper airways, rats were killed a
T.W.HESTERBERG ET AL. 105<br />
minimum <strong>of</strong> 24 h after the exposure was stopped; the right accessory lobe was<br />
frozen and later low temperature ashed for fiber recovery. Fiber lengths were<br />
determined using phase contrast optical microscopy, while fiber diameters were<br />
determined using scanning electron microscopy.<br />
Results from previous MMVF inhalation studies<br />
The results from two previous RCF inhalation studies (Davis et al., 1984;<br />
Smith et al., 1987) differ from the more recent RCC studies presented here.<br />
Davis et al., (1984) reported RCF exposure <strong>of</strong> rats resulted in an average <strong>of</strong><br />
5 per cent pulmonary fibrosis, pulmonary tumors in eight <strong>of</strong> 48 rats, and<br />
one peritoneal mesothelioma. The lower fibrosis and tumor response in the<br />
Davis study may have resulted from the lower exposure concentration used;<br />
8.4 mg m −3 compared to 30 mg m −3 in the present study. In addition, the<br />
use <strong>of</strong> fibers that were not presized, the use <strong>of</strong> whole-body exposure, or the<br />
fiber generation technique, which may have crushed some <strong>of</strong> the fibers,<br />
may account for the lack <strong>of</strong> consistency with the present study. Smith et<br />
al., (1987) exposed hamsters and rats to RCF at 200 f cm −3 , 6h a day, 5<br />
days a week, for 24 months. The rat study showed no significant increase<br />
in neoplasms and minimal pulmonary fibrosis in 22% <strong>of</strong> the exposed<br />
animals. In the hamster study, RCF produced only one mesothelioma in 50<br />
animals and no fibrosis was observed. It is difficult to explain why there<br />
was little response to RCF in the Smith studies, but it may be related to the<br />
different aerosol and exposure technology used or to the low exposure<br />
level; 12 mg m −3 compared to 30 mg m −3 in the present study.<br />
Previous inhalation studies <strong>of</strong> FG using rodents agree with the findings<br />
<strong>of</strong> the RCC studies. Fiber glass has been tested by inhalation in guinea pigs<br />
(Gross et al., 1970), hamsters (Lee et al., 1981; Smith et al., 1987), and<br />
rats (Gross et al., 1970; Lee et al., 1981; McConnell et al., 1984; Wagner<br />
et al., 1984; Mitchell et al. 1986; LeBouffant et al., 1987; Muhle et al.,<br />
1987; Smith et al., 1987). None <strong>of</strong> these studies identified a significant<br />
increase in either fibrosis or neoplasms following glass fiber inhalation in<br />
spite <strong>of</strong> FG lung burdens in excess <strong>of</strong> several hundred thousand fibers per<br />
mg dry lung tissue. In three <strong>of</strong> the above studies, the chronic inhalation<br />
toxicity <strong>of</strong> rock and slag wool were also examined (Wagner et al., 1984;<br />
LeBouffant et al., 1987; Smith et al., 1987). As was seen with fibrous glass,<br />
all three studies demonstrated no tumorigenic response by this route <strong>of</strong><br />
exposure.
106 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
Figure 7.2 (continued over) Lung burdens <strong>of</strong> (a) WHO fibers, (b) fibers >10 µm in<br />
length, and (c) fibers >20 µm in length per mg dry lung tissue from the lungs <strong>of</strong> rats<br />
continuously exposed to 30 mg m −3 <strong>of</strong> the five different MMVFs. Rats were killed<br />
at least 24 h after the exposure was stopped, the right accessory lobe was frozen<br />
and later low temperature ashed for fiber recovery.<br />
<strong>Industrial</strong> hygiene studies<br />
RCF<br />
<strong>Industrial</strong> hygiene monitoring data obtained on a regular basis at locations
Figure 7.2 Continued<br />
where RCF products are manufactured show that exposures are generally<br />
below 1.0 f cm −3 , typically below 0.2 f cm −3 . In a recent study, RCF levels<br />
during various end-user operations ranged from 0.12 to 1.55 f cm −3 with<br />
an overall mean and SD <strong>of</strong> 0.74±0.49 f cm −3 (Lees et al., 1993b). Other<br />
end-user studies have indicated that RCF exposures can exceed 5 f cm −3 or<br />
higher if appropriate engineering controls and work practices are not<br />
followed (Schuller, 1985–1988).<br />
Fiber glass<br />
T.W.HESTERBERG ET AL. 107<br />
Recently, studies which examined human aerosol exposure to fiber glass in<br />
manufacturing, installation and removal, and in ambient air were reviewed<br />
(Hesterberg and Hart, 1994). In most cases, human exposures to airborne<br />
fiber glass during manufacturing and installation fell well below the OSHAproposed<br />
permissible exposure limit (PEL) <strong>of</strong> 1 f cm −3 air (OSHA, 1992).<br />
Airborne fiber concentrations during FG manufacturing operations are<br />
typically less than 0.2 f cm −3 , with the majority being less than 0.1 f cm −3 .<br />
Exceptions include manufacture <strong>of</strong> finer diameter fiber glass and blowing<br />
installation <strong>of</strong> loose fiber glass that is either milled or lacks binder. Airborne<br />
levels averaging greater than 1 f cm −3 have been reported in the production<br />
<strong>of</strong> finer diameter fiber glass (TIMA, 1990), while blowing installation <strong>of</strong><br />
loose fiber glass without binder resulted in a task length average (TLA) <strong>of</strong><br />
7.67 f cm −3 , and an 8-h TWA <strong>of</strong> 1.96 f cm −3 (Lees et al., 1993a). Blowing<br />
installation <strong>of</strong> loose mineral wool also resulted in higher aerosol levels; a
108 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
TLA <strong>of</strong> 1.94 f cm −3 and an 8-h TWA <strong>of</strong> 0.97 f cm −3 (Lees et al., 1993a).<br />
Removal <strong>of</strong> fiber glass insulation created an aerosol <strong>of</strong> 0.042 f cm −3 (Jacob<br />
et al., 1993). Fiber concentrations <strong>of</strong> 0.004 f cm −3 were reported for<br />
buildings recently insulated with FG (Jacob et al., 1992). However this figure<br />
includes all types <strong>of</strong> fibers as it was obtained using optical microscopy. The<br />
background level prior to fiber glass installation was 0.001 f cm −3 .<br />
Ambient environmental exposures to airborne vitreous fibers were<br />
extremely low; exposure levels <strong>of</strong> product-related vitreous fibers reported<br />
for outdoor air was 0.0007 f cm −3 (Tiesler and Draeger, 1994).<br />
In addition to manufacturing and field use surveys, release <strong>of</strong> fibrous<br />
glass during actual use <strong>of</strong> products, particularly fiber released from air<br />
filter media, has been monitored. To determine possible exposure <strong>of</strong><br />
building occupants to fibrous glass, ambient air was sampled in a number<br />
<strong>of</strong> public buildings in which fibrous glass air filtration products had been<br />
installed. These evaluations showed no significant release <strong>of</strong> fibers from the<br />
filters (Balzer et al., 1971; Cholak and Schafer, 1971).<br />
To evaluate the efficiency <strong>of</strong> fibrous glass filter blankets, several high<br />
volume air samples were collected at various points in the ductwork <strong>of</strong> a<br />
large <strong>of</strong>fice complex at the intake and the exhaust prior to changing the<br />
filter media, and at the exhaust 23 days after installation <strong>of</strong> the new filter.<br />
Analyses <strong>of</strong> the samples using electron microscopy indicate little initial<br />
fiber release which decreases rapidly thereafter to the limit <strong>of</strong> detection<br />
(Schuller, 1987).<br />
Rock and slag wool<br />
Airborne concentrations <strong>of</strong> dust and fibers reported from US mineral wool<br />
plants is generally higher than in US glass wool facilities. This includes both<br />
airborne fibers and total particulate matter. Fiber levels reported ranged<br />
from 0.01 to 1.4 f cm −3 , compared with 0.1–0.3 f cm −3 for glass wool.<br />
Total particulate matter sample results ranged from 0.05 to 23.6 mg m −3 in<br />
the mineral wool facilities and 0.09–8.48 mg m −3 for glass wool (Esmen et<br />
al., 1980).<br />
Comparison <strong>of</strong> Human MMVF exposures used in the<br />
recent rat chronic inhalation studies<br />
When using animal inhalation studies for assessment <strong>of</strong> potential risk to<br />
human health <strong>of</strong> airborne fibers, it is critical to demonstrate that the<br />
characteristics and concentrations <strong>of</strong> the experimental fiber aerosols are<br />
comparable with those in human exposure situations. It is also important<br />
for risk assessment that the actual target organ dose in the animal model<br />
reach or exceed that found in exposed humans. To illustrate, consider<br />
levels <strong>of</strong> fiber glass published in a number <strong>of</strong> recent reports. A qualitative
T.W.HESTERBERG ET AL. 109<br />
Table 7.6 Representative airborne levels <strong>of</strong> fiber glass in workplace and rat<br />
inhalation study<br />
a Outdoor data from Tiesler and Draeger (1994). Product-related fibers counted<br />
using NIOSH A Rules.<br />
b Data from Jacob et al., (1993). Airborne levels resulting from manufacturing<br />
operations using FG insulation.<br />
c Data from Lees et al., (1993a). Installation <strong>of</strong> residential insulation.<br />
d Jacob et al. (1992) reported that levels returned to background within hours after<br />
Batt installation.<br />
All other data are averages from the various studies herein cited.<br />
and quantitative comparison was made <strong>of</strong> the aerosol and lung fibers in the<br />
rat inhalation study with those in various human exposure situations<br />
(Hesterberg and Hart, 1994). A comparison <strong>of</strong> the reported aerosol fiber<br />
levels in various human settings with those used in the rat inhalation study<br />
is shown in Table 7.6. FG levels in the rat aerosol were more than five<br />
orders <strong>of</strong> magnitude higher than the reported level for outdoor air, and at<br />
least three orders <strong>of</strong> magnitude higher than for average airborne levels for<br />
many occupational settings (e.g. over 2000fold higher than FG batt<br />
installation). The rat aerosol was 75-fold more concentrated than the<br />
highest reported average TWA for airborne fiber levels in an occupational<br />
setting, i.e. blowing installation <strong>of</strong> unbound fiber glass (the potential for<br />
higher airborne levels has been recognized for some time, and<br />
recommended work practices call for the use <strong>of</strong> respirators in such<br />
circumstances). Despite the range in products and occupational settings,<br />
fiber dimensions in most <strong>of</strong> the human exposures examined were fairly<br />
similar to those found in the rat inhalation study aerosol (Hesterberg and<br />
Hart, 1994). The fiber dimensions <strong>of</strong> aerosolized rock and slag wool<br />
collected from workplace air during the installation <strong>of</strong> batts or blowing <strong>of</strong><br />
loose fibers have similar mean diameters to that <strong>of</strong> fiber glass (1.0–1.6<br />
µm). However, the mean lengths appear to be greater (30–50 µm) than for<br />
most workplace samples <strong>of</strong> fiber glass.<br />
Hesterberg and Hart (1994) also compared the lung burdens <strong>of</strong> rats<br />
exposed in the recent fiber glass inhalation study in rats with lung burdens<br />
found in workers involved in MMMF (primarily FG) manufacturing<br />
(McDonald et al. 1990). As shown in Table 7.7, rat fiber glass lung<br />
burdens vastly exceeded that <strong>of</strong> the workers reported by McDonald et al.,
110 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
which was not significantly elevated above reference levels. Fibers per mg<br />
dry lung for the rat after lifetime exposure was >4000-fold higher than for<br />
the fiber glass worker, average exposure 11 years (the average time from<br />
last employment in MMMF manufacturing and death was 12 years). Lung<br />
fiber dimensions in the rat study were comparable to those <strong>of</strong> fibers<br />
recovered from the lung tissue <strong>of</strong> fiber glass manufacturing workers. From<br />
these comparisons, it can be concluded that the exposure levels used in the<br />
recent rat inhalation studies unequivocally achieved the goal <strong>of</strong> the studies<br />
to exceed human exposures by several orders <strong>of</strong> magnitude.<br />
Summary and conclusions<br />
MMVFs are among the most studied commercial products due to their<br />
widespread use and the concern for potential health effects <strong>of</strong> respirable<br />
fibers. In recent animal inhalation studies RCF produced lung fibrosis,<br />
mesotheliomas, and significant increases in lung tumors. However, it is<br />
believed that any potential cancer risk from RCF exposure can be<br />
minimized, if not eliminated, because <strong>of</strong> the small number <strong>of</strong> workers<br />
exposed and the ability to use respiratory protection and engineering<br />
controls to limit worker exposure. Both human and animal inhalation<br />
studies have shown no association between fiber glass exposure and<br />
disease. Although high exposure levels <strong>of</strong> rock wool (several orders <strong>of</strong><br />
magnitude higher than most reported workplace exposures) produced<br />
minimal lung fibrosis in rats, no mesotheliomas and no significant increase<br />
in lung tumors were observed. Slag wool produced no fibrosis or increase<br />
in tumors in the animal studies. The cohort mortality studies <strong>of</strong> rock wool<br />
and slag wool workers have also provided no clear dose-response<br />
relationship with fiber exposure.<br />
Results from the combined animal inhalation studies showed that<br />
differences in lung fiber burdens and lung clearance rates could not explain<br />
the differences observed in the toxicologic effects <strong>of</strong> MMVFs. These<br />
findings clearly indicate that dose, dimension and durability (i.e. the<br />
persistence <strong>of</strong> fibers in the rat lung) are not the only determinants <strong>of</strong> fiber<br />
toxicity; chemical composition and the surface physicochemical properties<br />
<strong>of</strong> the fibers may also play an important role. Exposure levels from animal<br />
inhalation studies were at least three orders <strong>of</strong> magnitude higher than for<br />
average airborne levels reported for many occupational settings.
Table 7.7 Reported lung fiber levels from fiber glass workers and rat inhalation study<br />
T.W.HESTERBERG ET AL. 111<br />
a Lung fibers: for humans, NIOSH A rules; for rats, total fibers (all fibers length/diameter >3:1).<br />
b For humans, NIOSH A rules; for rats, WHO respirable fibers, comparable to A rules because there were no diameters >3 µm in<br />
rats.<br />
c Hesterberg et al., (1993b), Rat fiber exposure was 5 days week•1 , 6 h day•1 for lifetime (2 years).<br />
d McDonald et al., (1990). Negative controls had not worked with FG and were matched with each FG worker for age and<br />
locale.<br />
e Occupational exposures averaged 11 years, followed by average <strong>of</strong> 12 years without exposure prior to death.<br />
f 101 were FG workers; 11 were mineral wool workers.<br />
g Not reported.
112 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS<br />
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8<br />
Pulmonary Toxicity Studies with Man-made<br />
Organic Fibres: Preparation and Comparisons <strong>of</strong><br />
Size-separated Para-aramid with Chrysotile<br />
Asbestos Fibres<br />
DAVID B.WARHEIT, 1 MARK A.HARTSKY, 1 CHARLES<br />
J.BUTTERICK 2 and STEVEN R.FRAME 1<br />
1 DuPont Haskell Laboratory, Newark, DE, 2 Texas Tech<br />
Health Sciences Center Lubbock, TX<br />
Introduction<br />
This study was designed to compare the pulmonary toxic effects <strong>of</strong><br />
inhaled, size-separated preparations <strong>of</strong> chrysotile asbestos fibres with paraaramid<br />
fibrils at similar aerosol fibre concentrations. Chrysotile samples<br />
are known to have an abundance <strong>of</strong> short fibres, with mean lengths<br />
generally in the range <strong>of</strong> 2 µm. This is important to note because one <strong>of</strong> the<br />
critical factors influencing the pathogenesis <strong>of</strong> fibre-related lung disease is<br />
fibre dimension (Davis et al., 1986). As a consequence, attempts were made<br />
to selectively enhance the mean lengths <strong>of</strong> chrysotile asbestos fibres used in<br />
this inhalation toxicity study, in order to make reasonable comparisons<br />
between the two fibre-types.<br />
Methods<br />
General experimental design<br />
Groups <strong>of</strong> male Crl: CDBR rats (7–8 weeks old, Charles River Breeding<br />
Laboratories, Kingston, New York) were used to assess the pulmonary<br />
effects <strong>of</strong> 2-week inhalation exposures to size-separated preparations <strong>of</strong><br />
Kevlar ® para-aramid fibrils or chrysotile asbestos fibres. Animals were<br />
exposed 6 hr day −1 , 5 days week −1 for 2 weeks. For this study, Kevlar ® was<br />
utilized as a representative para-aramid fibril. The two commercial types <strong>of</strong><br />
para-aramid fibres are Twaron ® , made by Akzo, and Kevlar ® , made by<br />
DuPont. Following exposure, the lungs <strong>of</strong> p-aramid or chrysotile-exposed<br />
animals and age-matched sham controls were subsequently evaluated by<br />
bronchoalveolar lavage fluid analysis at 0 h, 5 days, 1 and 3 months<br />
postexposure. The lungs <strong>of</strong> additional animals were evaluated for<br />
biodurability, pulmonary clearance, pulmonary histopathologic lesions and
118 PULMONARY TOXICITY STUDIES WITH MAN-MADE ORGANIC FIBRES<br />
lung and mesothelial cell proliferation at 0 hrs, 5 days 1, 3, 6 and 12<br />
months postexposure.<br />
Fibre preparation and inhalation exposure<br />
Ultrafine Kevlar ® p-aramid fibrils were supplied by DuPont Fibres. A<br />
special preparation <strong>of</strong> respirable p-aramid fibrils which had been prepared<br />
for the 2-year inhalation study (Lee et al., 1988) was utilized for this study.<br />
Bulk Canadian chrysotile asbestos fibres were obtained from Mr John<br />
Addison <strong>of</strong> the Institute <strong>of</strong> Occupational Medicine in Edinburgh, Scotland.<br />
Attempts were made to size-separate the bulk fibre preparation (i.e.<br />
selectively enhance the percentages <strong>of</strong> long fibres while removing the short<br />
fibres) by placing the fibres in a rotating sieve shaker and sieving through a<br />
series <strong>of</strong> metal mesh screens. The fraction containing the longer fibres (and<br />
a number <strong>of</strong> short fibres) was collected and generated for inhalation<br />
studies; fibres were collected on a filter and dimensional analysis (i.e. length<br />
and diameter assessments) was performed using scanning electron<br />
microscopy. The results showed that this technique was partially successful<br />
as the median and mean lengths <strong>of</strong> fibres were increased from 3 and 5 µm,<br />
respectively, in the original bulk sample to 6 and 9 µm in the generated<br />
sample preparation. The median lengths and diameters <strong>of</strong> p-aramid fibrils<br />
used in the study were 9 µm and 0.3 µm, respectively.<br />
The methods for aerosol generation <strong>of</strong> p-aramid fibrils have previously<br />
been reported (Warheit et al., 1992). Final mean fibre concentrations for<br />
the p-aramid exposures were 772 and 419 f cm −3 .<br />
Aerosols <strong>of</strong> chrysotile asbestos fibres were generated in a similar<br />
manner, i.e. with a binfeeder and baffles, but without the microjet<br />
apparatus. Final mean fibre concentrations for the chrysotile asbestos<br />
exposures were 782 and 458 f cm −3 . Fibre lung burdens were quantified<br />
from digested lung tissue <strong>of</strong> animals sacrificed immediately after the end <strong>of</strong><br />
the 2-week exposure.<br />
Pulmonary lavage and biochemical assays on lavaged<br />
fluids<br />
Bronchoalveolar lavage procedures, cell quantification, and biochemical<br />
assays were conducted according to methods previously described (Warheit<br />
et al., 1984a, 1992). In addition, the methods for measuring lactate<br />
dehydrogenase (LDH), N-acetyl-β-glucosaminidase (NAG), and alkaline<br />
phosphatase (ALP) and protein in BAL fluids have been reported (Warheit<br />
et al., 1992).
Lung dissection, tissue preparation and pulmonary cell<br />
proliferation<br />
The lungs <strong>of</strong> rats exposed to p-aramid and chrysotile asbestos fibres for 2<br />
weeks were prepared for light microscopy by airway infusion using<br />
methods previously reported (Warheit et al., 1984b, 1991).<br />
Pulmonary cell proliferation experiments were designed to measure the<br />
effects <strong>of</strong> fibre inhalation exposure on terminal bronchiolar, proximal lung<br />
parenchymal (i.e. alveolar duct bifurcations and adjacent areas), subpleural<br />
and visceral pleural, and mesothelial cell turnover in rats following 2-week<br />
exposures. Groups <strong>of</strong> sham and fibre-exposed rats were given a 2-h pulse<br />
immediately after exposures, as well as 5 days, 1, 3, 6 and 12 months (still<br />
in progress) postexposure with an intraperitoneal injection <strong>of</strong> 5-bromo-2′deoxy-uridine<br />
(BrdU) dissolved in a 0.5N sodium bicarbonate buffer<br />
solution at a dose <strong>of</strong> 100 mg kg −1 body weight as previously described<br />
(Warheit et al., 1992). In addition, sections <strong>of</strong> duodenum served as a<br />
positive control. For each treatment group, there were immunostained<br />
nuclei in airways (i.e. terminal bronchiolar epithelial cells), lung parenchyma<br />
(i.e. epithelial, interstitial cells or macrophages), subpleura and visceral<br />
pleura, and mesothelial cells. All regions were counted by light microscopy<br />
at ×1000 magnification. Statistics were carried out using a two-tailed<br />
Students t test on a Micros<strong>of</strong>t Excel s<strong>of</strong>tware program.<br />
Fibre recovery from lung tissue<br />
Para-aramid fibrils were recovered from the lungs <strong>of</strong> exposed rats using a<br />
diluted 1.3% hypochlorite (Clorox bleach) solution. The results <strong>of</strong><br />
validation studies in our laboratory demonstrated that the dilute Clorox<br />
solution (10 min digestion) was more effective in digesting lung tissue than<br />
the KOH method that we had previously reported (Warheit et al., 1992).<br />
Chrysotile asbestos fibres were recovered from the lungs <strong>of</strong> exposed rats<br />
by incubating the lung tissue with a 5.25% hypochlorite solution for 3 h.<br />
Subsequently, the filters containing fibres recovered from lung tissue were<br />
mounted and prepared for phase-contrast light microscopy (for counting)<br />
and for scanning electron microscopy (for fibre dimensional analysis),<br />
according to methods previously described (Warheit et al., 1992).<br />
Results<br />
D.B.WARHEIT ET AL. 119<br />
Size-separation methods for chrysotile asbestos fibres<br />
The results from size-separation attempts showed that there was a shift in<br />
the distribution <strong>of</strong> fibre lengths from shorter fibres to longer fibres<br />
(Figures 8.1(A)– (C)). Count median lengths <strong>of</strong> chrysotile asbestos fibres
120 PULMONARY TOXICITY STUDIES WITH MAN-MADE ORGANIC FIBRES<br />
were increased from 3 µm in the original generated sample to 6 µm in the<br />
size-separated sample. In comparison to the chrysotile asbestos sample,<br />
there was a significantly greater proportion <strong>of</strong> long p-aramid fibrils which<br />
were used in the inhalation study with median lengths >9 µm.<br />
Lung burden analysis<br />
Although the aerosol fibre concentrations were similar throughout the<br />
study (p-aramid high conc.=772 f cm −3 , chrysotile high conc.=782 f cm −3 ;<br />
p-aramid low conc.=419 f cm −3 , chrysotile low conc.=458 f cm −3 ),<br />
measurement <strong>of</strong> lung fibre burdens from digested lung tissue at time 0 (i.e.<br />
immediately after exposure) demonstrated a substantial difference in lung<br />
burden between the two fibre-types as measured by phase contrast optical<br />
microscopy (PCOM). The mean lung fibre (>5 µm) burden from 3 rats/<br />
dose group exposed to chrysotile asbestos was 3.7×10 7 (±7.4×10 6 ) fibres/<br />
lung for the high dose group and 1.3×10 7 (±4×10 6 ) fibres/lung for the low<br />
dose group. In contrast, the mean lung fibre burden from 3 rats/dose group<br />
exposed to para-aramid fibres was 7.6×10 7 (±1.9×10 7 ) fibres per lung for<br />
the high dose group and 4.8×10 7 ( ±2.1×10 7 ) fibres/lung for the low dose<br />
group. In addition, the count median length <strong>of</strong> chrysotile fibres recovered<br />
from the lungs <strong>of</strong> exposed animals immediately after 2-week exposure was<br />
3.5 µm, while the count median diameter was 0.15 µm. In contrast, the<br />
count median length <strong>of</strong> para-aramid fibres recovered from the lungs <strong>of</strong><br />
exposed animals immediately after 2-week exposure was 8.6 µm, while the<br />
count median diameter was 0.3 µm (Figure 8.2(A) and (B); numerical data<br />
not shown). These data indicate that our attempts to size-separate<br />
Canadian chrysotile fibres were only partially successful. The lung burden<br />
data also suggest that comparisons <strong>of</strong> the effects <strong>of</strong> chrysotile vs paraaramid<br />
at high and low doses are difficult to make since the doses were not<br />
equivalent.<br />
Bronchoalveolar lavage data<br />
Two-week exposures to p-aramid fibrils or chrysotile asbestos fibres<br />
produced transient pulmonary inflammatory responses as measured by<br />
bronchoalveolar lavage fluid analysis (see Table 8.1).<br />
Light microscopic histopathology<br />
Exposures to p-aramid and chrysotile were associated with minimal to mild<br />
centriacinar inflammation and fibrosis (increased trichrome staining)<br />
immediately after and 5 days after 2-week exposures. Lesions were slightly<br />
more prominent in p-aramid-exposed rats due to increased inflammation.<br />
Lesions were less severe at 1 month and essentially resolved at 6 months
D.B.WARHEIT ET AL. 121<br />
Figure 8.1 (A) Chrysotile asbestos lengths—original generated sample for 4 different<br />
experiments. The graph depicts the fibre length distributions as assessed by<br />
scanning electron microscopy from four aerosol exposures prior to attempts to size<br />
separate the fibres. Fifty percent <strong>of</strong> the fibres from all four groups are less than 3–4<br />
µm. (B) Distributions <strong>of</strong> size-separated chrysotile asbestos lengths used in the<br />
inhalation study from the high-dose
122 PULMONARY TOXICITY STUDIES WITH MAN-MADE ORGANIC FIBRES<br />
Figure 8.1 Continued<br />
exposures and (C) the low-dose exposure groups. A casual glance at the two graphs<br />
B and C indicates that some success was attained in increasing the mean lengths in<br />
the aerosol <strong>of</strong> the generated chrysotile asbestos sample.<br />
with only occasional centriacinar regions having slight, fibril-associated<br />
thickening <strong>of</strong> alveolar duct bifurcations. At 1 year postexposure, the lungs<br />
in p-aramid exposed rats were similar to controls. The 1-year chrysotileexposed<br />
animals are still in recovery.<br />
Pulmonary cell proliferation<br />
In chrysotile asbestos-exposed rats, substantial increases compared to<br />
controls in pulmonary cell proliferation indices were measured on terminal<br />
bronchiolar, parenchymal, visceral pleural/subpleural and mesothelial<br />
surfaces, and many <strong>of</strong> these effects were sustained through 3 months<br />
postexposure. These data demonstrate that 2-week chrysotile exposures<br />
produced a prolonged proliferative response in airway, alveolar and<br />
subpleural cells, as evidenced by the sustained effect through 3 months<br />
postexposure (Table 8.2).<br />
Pulmonary cell proliferation studies demonstrated that 2-week exposures<br />
to the high dose <strong>of</strong> p-aramid fibrils produced a transient increase in<br />
terminal bronchiolar and visceral pleural/subpleural cell labeling responses.<br />
No increases in lung parenchymal, or subpleural cell labeling indices were<br />
mea sured at any time period relative to sham controls. In addition, no
D.B.WARHEIT ET AL. 123<br />
Figure 8.2 (A) Scanning electron microscopy (SEM) micrograph <strong>of</strong> an aerosol filter<br />
containing a mixture <strong>of</strong> long and short chrysotile asbestos fibres (arrows). (B) An<br />
SEM micrograph <strong>of</strong> fibres recovered from the lung <strong>of</strong> a rat 3 months after 2-week<br />
chrysotile exposures. Note that most <strong>of</strong> the fibres are long (arrows), indicating that<br />
the long chrysotile asbestos fibres were retained in the lung while the shorter fibres<br />
were cleared from the respiratory tract.<br />
increases in cell labeling indices were measured in animals exposed to a<br />
lower dose <strong>of</strong> p-aramid fibrils at any postexposure time period (Table 8.2).
124 PULMONARY TOXICITY STUDIES WITH MAN-MADE ORGANIC FIBRES<br />
Table 8.1 Pulmonary inflammation and fibre biodurability in the lungs <strong>of</strong><br />
chrysotile asbestos and p-aramid-exposed rats<br />
0 h=immediately after exposure; 5 D=5 days; 1 M=1 month; 3 M=3 months; 6<br />
M=6 months;<br />
ND=not determined.<br />
Lung digestion/biodurability studies<br />
Preliminary dimensional analysis studies demonstrated that median lengths<br />
<strong>of</strong> fibres recovered from digested asbestos-exposed lung tissue were<br />
increased over time suggesting that short asbestos fibres were selectively<br />
cleared from the lungs, with apparent insignificant or pulmonary clearance<br />
and greater durability/retention <strong>of</strong> long fibres (Table 8.1).<br />
Preliminary studies with p-aramid fibrils recovered from the lungs <strong>of</strong><br />
exposed rats are consistent with earlier data suggesting biodegradability <strong>of</strong><br />
inhaled p-aramid fibrils (Warheit et al., 1992; Kelly et al., 1993)<br />
(Table 8.1). These data also are in agreement with the results <strong>of</strong> a current<br />
interim report authored by the Institute <strong>of</strong> Occupational Medicine in<br />
Edinburgh, Scotland. In addition, as previously reported (Warheit et al.,<br />
1992), a transient increase in fibre numbers at early postexposure time<br />
periods was measured following cessation <strong>of</strong> exposure. These results<br />
indicate that the increase in p-aramid fibres is due to fibre shortening and as<br />
a consequence, increased numbers <strong>of</strong> shorter fibres. This is accounted for<br />
by a substantial reduction in the median lengths <strong>of</strong> recovered fibres<br />
concomitant with only a slight decrease in fibre diameter.
Table 8.2 Cell proliferation effects in chrysotile asbestos and p-Aramid-exposed<br />
rats<br />
a p
126 PULMONARY TOXICITY STUDIES WITH MAN-MADE ORGANIC FIBRES<br />
The BrdU pulmonary cell labeling results demonstrating sustained<br />
proliferative effects in chrysotile-exposed rats presented here are consistent<br />
with findings from several other investigators (Brody and Overby, 1989;<br />
McGavran et al., 1990; Coin et al., 1992a). In studies by Brody and<br />
Overby (1989), acute inhalation exposures to chrysotile asbestos fibres<br />
produced a biphasic cell labeling response in the lungs <strong>of</strong> exposed rats and<br />
mice. This was characterized by dramatic increases in epithelial cell DNA<br />
synthesis, followed several days later by enhanced labeling <strong>of</strong> interstitial<br />
cells. In follow-up studies, a 3 day exposure prolonged the duration <strong>of</strong><br />
increased cell labeling (Coin et al., 1992b). In another study, Coin et al.,<br />
(1991) reported that a 5-h exposure to chrysotile fibres in mice produced<br />
substantial increases in mesothelial and subpleural cell labeling indices at 2<br />
and 8 days postexposure.<br />
The finding <strong>of</strong> sustained subpleural and mesothelial cell proliferation in<br />
chrysotile-exposed rats was unexpected and raises the issue regarding the<br />
association <strong>of</strong> chrysotile with the development <strong>of</strong> mesothelioma. In this<br />
regard, inhalation <strong>of</strong> chrysotile asbestos fibres produced mesotheliomas in<br />
exposed rats (Wagner et al., 1974; Davis and Jones, 1988).<br />
The biodurability data reported here demonstrating retention or reduced<br />
clearance <strong>of</strong> long chrysotile fibres are consistent with the results <strong>of</strong><br />
previous studies by Roggli and Brody (1984) and Bellmann et al., (1986,<br />
1987). In contrast to the enhanced biodurability <strong>of</strong> chrysotile asbestos<br />
fibres, the results with p-aramid fibres suggest that the fibrils undergo<br />
biodegradability in the lungs <strong>of</strong> exposed rats. These findings confirm our<br />
earlier studies (Warheit et al., 1992) and are in concordance with the<br />
results <strong>of</strong> Kelly et al. (1993) and the recent findings <strong>of</strong> the IOM.<br />
In conclusion, size separation techniques for chrysotile asbestos fibres<br />
were partially successful in increasing median lengths from 3 µm to 6 µm.<br />
Histopathological studies demonstrated that both p-aramid and chrysotile<br />
produced a minimal to mild inflammatory response which produced<br />
thickening <strong>of</strong> the alveolar duct bifurcations. These effects peaked at 1<br />
month postexposure and were essentially reversible by 6 months<br />
postexposure.<br />
Pulmonary cell labeling studies demonstrated substantial increases in<br />
lung parenchymal, airway, pleural/subpleural, and mesothelial cell<br />
proliferation effects following chrysotile exposures, suggesting that<br />
chrysotile produces a potent proliferative response in the airways, lung<br />
parenchyma, and subpleural/ pleural regions. In contrast, p-aramid<br />
exposures produced only transient effects in airway and subpleural regions.<br />
Fibre biopersistence/durability results thus far indicate that the long<br />
chrysotile fibres are retained in the lung or cleared at a slow rate. In<br />
contrast, p-aramid fibres have low biodurability in the lungs <strong>of</strong> exposed<br />
animals. In this regard, median lengths <strong>of</strong> chrysotile fibres recovered from
exposed lung tissue were increased over time, while median lengths <strong>of</strong> paramid<br />
fibrils were decreased over time.<br />
It is concluded that the proliferative effects and enhanced biodurability<br />
<strong>of</strong> chrysotile that are associated with the induction <strong>of</strong> chronic disease do<br />
not occur with p-aramid fibrils. Therefore, inhalation <strong>of</strong> chrysotile asbestos<br />
fibres is likely to produce greater long-term pulmonary toxic effects in<br />
comparison to para-aramid fibrils.<br />
Acknowledgments<br />
This study was sponsored by the DuPont Co. and Akzo Nobel Corp.<br />
References<br />
D.B.WARHEIT ET AL. 127<br />
BELLMANN, B., KONIG, H., MUHLE, H. and POTT, F., 1986, Chemical<br />
durability <strong>of</strong> asbestos and <strong>of</strong> man-made mineral fibres in vivo, J. Aerosol Sci.,<br />
17, 341–5.<br />
BELLMANN, B., MUHLE, H., POTT, F., KONIG, H., KLOPPEL, H. and<br />
SPURNY, K., 1987, Persistence <strong>of</strong> man-made mineral fibers (MMMF) and<br />
asbestos in rat lungs, Ann. Occup. Hyg., 31(4B), 693–709.<br />
BRODY, A.R. and OVERBY, L.H., 1989, Incorporation <strong>of</strong> tritiated thymidine by<br />
epithelial and interstitial cells in bronchiolar-alveolar regions <strong>of</strong> asbestosexposed<br />
rats, Am. J. Pathol., 134, 133–40.<br />
COIN, P.G., MOORE, L.B., ROGGLI, V. and BRODY, A.R., 1991, Pleural<br />
incorporation <strong>of</strong> 3 H-TdR after inhalation <strong>of</strong> chrysotile asbestos in the mouse,<br />
Am. Rev. Respir. Dis., 143, A604.<br />
COIN, P.G., ROGGLI, V.L. and BRODY, A.R., 1992a, Deposition, clearance and<br />
translocation <strong>of</strong> chrysotile asbestos from peripheral and central regions <strong>of</strong> the<br />
rat lung, Environ, Res., 58, 97–116.<br />
COIN, P.G., ROGGLI, V. and BRODY, A.R., 1992b, Pulmonary fibrogenesis and<br />
BRDU incorporation after three consecutive inhalation exposures to chrysotile<br />
asbestos, Am. Rev. Respir. Dis., 145, A328.<br />
DAVIS, J.M.G. and JONES, A.D., 1988, Comparisons <strong>of</strong> the pathogenicity <strong>of</strong> long<br />
and short fibres <strong>of</strong> chrysotile asbestos in rats, Br. J. Exp. Pathol., 69, 717–37.<br />
DAVIS, J.M.G., ADDISON, J., BOLTON, R.E., DONALDSON, K. et al., 1986,<br />
The pathogenicity <strong>of</strong> long versus short fibre samples <strong>of</strong> amosite asbestos<br />
administered to rats by inhalation or intraperitoneal injection, Br. J. Exp.<br />
Pathol, 67, 415–30.<br />
KELLY, D.P., MERRIMAN, E.A., KENNEDY, G.L.JR. and LEE, K.P., 1993,<br />
Deposition, clearance, and shortening <strong>of</strong> Kevlar para-aramid fibrils in acute,<br />
subchronic, and chronic inhalation studies in rats, Fundam. Appl. Toxicol, 21,<br />
345–54.<br />
LEE, K.P., KELLY, D.P., O’NEAL, F.O., STADLER, J.C. and KENNEDY, G.<br />
L.JR, 1988, Lung response to ultrafine Kevlar aramid synthetic fibrils<br />
following 2-year inhalation exposure in rats, Fundam. Appl. Toxicol., 11, 1–<br />
20.
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McGAVRAN, P.D., BUTTERICK, C.J. and BRODY, A.R., 1990, Tritiated<br />
thymidine incorporation and the development <strong>of</strong> an interstitial lesion in the<br />
bronchiolar alveolar regions <strong>of</strong> the lungs <strong>of</strong> normal and complement deficient<br />
mice after inhalation <strong>of</strong> chrysotile asbestos, J. Environ. Pathol. Toxicol.<br />
Oncol, 9, 377–92.<br />
ROGGLI, V.L. and BRODY, A.R., 1984, Changes in numbers and dimensions <strong>of</strong><br />
chrysotile asbestos fibers in lungs <strong>of</strong> rats following short-term exposure, Exp.<br />
Lung Res., 7, 133–47.<br />
WAGNER, J.C., BERRY, G., SKIDMORE, J.W. and TIMBRELL, V., 1974, The<br />
effects <strong>of</strong> the inhalation <strong>of</strong> asbestos in rats, Br. J. Cancer, 29, 252–70.<br />
WARHEIT, D.B., HILL, L.H. and BRODY, A.R., 1984a, Surface morphology and<br />
correlated phagocytic capacity <strong>of</strong> pulmonary macrophages lavaged from the<br />
lungs <strong>of</strong> rats, Exp. Lung Res., 6, 71–82.<br />
WARHEIT, D.B., CHANG, L.Y., HILL, L.H., HOOK, G.E.R., CRAPO, J.D. and<br />
BRODY, A.R., 1984b, Pulmonary macrophage accumulation and<br />
asbestosinduced lesions at sites <strong>of</strong> fiber deposition, Am. Rev. Respir. Dis., 129,<br />
301.<br />
WARHEIT, D.B., CARAKOSTAS, M.C., HARTSKY, M.A. and HANSEN, J.F.,<br />
1991, Development <strong>of</strong> a short-term inhalation bioassay to assess pulmonary<br />
toxicity <strong>of</strong> inhaled particles: Comparisons <strong>of</strong> pulmonary responses to carbonyl<br />
iron and silica, Toxicol Appl. Pharmacol., 107, 350–68.<br />
WARHEIT, D.B., KELLAR, K.A. and HARTSKY, M.A., 1992, Pulmonary cellular<br />
effects in rats following aerosol exposures to ultrafine Kevlar® aramid fibrils:<br />
evidence for biodegradability <strong>of</strong> inhaled fibrils, Toxicol. Appl. Pharmacol,<br />
116, 225– 39.
9<br />
Pulmonary Hyperreactivity to <strong>Industrial</strong><br />
Pollutants<br />
JÜRGEN PAULUHN<br />
Bayer AG, Wuppertal<br />
Introduction<br />
Environmental agents, such as ozone, nitrogen dioxide, formaldehyde, and<br />
sulfur dioxide; occupational pollutants, including natural dusts (grain, red<br />
cedar, animal dander), irritant fumes or vapors, and organic acid<br />
anhydrides, reactive dyes, or (di)isocyanates can cause increases in airway<br />
reactivity. Airway hyperreactivity is defined as an exaggerated acute<br />
obstructive response <strong>of</strong> the airways to one or more nonspecific stimuli. The<br />
incriminated etiologic low-molecular-weight agents all share a common<br />
toxicological characteristic <strong>of</strong> being irritant in nature. In some cases, the<br />
agents are present as a gas, in others the inciting agent is an aerosol. As yet<br />
it is not clear, for instance, whether induced airway hyperreactivity is a<br />
dose-effect phenomenon and whether a brief high level exposure causes<br />
more prolonged or intense airways response. While the illness clinically<br />
simulates bronchial asthma and is associated with airway hyperreactivity,<br />
it is considered to be different from typical occupational asthma because <strong>of</strong><br />
its rapid onset, specific relationship to a single environmental exposure,<br />
and no apparent preexisting period <strong>of</strong> sensitization to occur with the<br />
apparent lack <strong>of</strong> an allergic or immunologic etiology. Hence, this illness is<br />
termed reactive airways dysfunction syndrome, or RADS, because the<br />
characteristic finding is hyperreactivity <strong>of</strong> the airways (Brooks et al.,<br />
1985). Mechanisms to explain the development <strong>of</strong> RADS focus on the<br />
toxic effects <strong>of</strong> the irritant exposure on the airways. How this increased<br />
bronchial responsiveness is precisely triggered, amplified, sustained and<br />
how it relates to inflammatory events remains, to a certain extent,<br />
incompletely elucidated (Kay, 1991).<br />
A common pathologie accompaniment or cause <strong>of</strong> increased airway<br />
hyper-responsiveness is pulmonary inflammation. It is suggested that this<br />
inflammation is responsible for the change in histamine or cholinergic<br />
agonist responsiveness. Because subepithelial irritant receptors are<br />
superficial in location, they could be affected by an extensive bronchial<br />
inflammatory response which might occur after heavy irritant exposure.
130 PULMONARY HYPERREACTIVITY TO INDUSTRIAL POLLUTANTS<br />
Subsequent re-epithelialization and probable reinervation <strong>of</strong> bronchial<br />
mucosa might drastically alter the threshold <strong>of</strong> the receptors and cause<br />
airways hyperreactivity. It has been hypothesized that damage to airway<br />
epithelium by irritant chemicals could decrease the threshold <strong>of</strong> sensory<br />
endings within the mucosa, resulting in increased afferent and efferent vagal<br />
activity. Airway mucosal inflammation, activation <strong>of</strong> airway afferent<br />
nerves, and the release <strong>of</strong> low-molecular-weight neuropeptides as<br />
mediators <strong>of</strong> inflammation are known to affect the tonus <strong>of</strong> the airway<br />
smooth muscle and may play a crucial role in the acute increase in airway<br />
hyperresponsiveness occurring after exposure to irritant or inflammatory<br />
stimuli. Additionally, inflammatory mediators may further attract and<br />
activate inflammatory cells, which themselves release a whole array <strong>of</strong><br />
chemotactic and cytotoxic mediators that serve to perpetuate and amplify<br />
the inflammatory response. This complex interaction <strong>of</strong> different factors<br />
may result in epithelial desquamation, mucus gland hyperplasia, smooth<br />
muscle hypertrophy, and eventually render the airways hyperreactive to<br />
specific as well as nonspecific stimuli.<br />
Increased bronchial irritability, or hyperresponsiveness, to a wide variety<br />
<strong>of</strong> chemical agents and physical stimuli is also a major characteristic<br />
feature <strong>of</strong> bronchial asthma and the reactive airways dysfunction syndrome<br />
might clinically be indistinguishable from the asthma syndrome. Also for<br />
the latter, particular attention has been placed on the role <strong>of</strong> inflammation<br />
mediated influx <strong>of</strong> cells, mediator release and the interaction <strong>of</strong> irritant<br />
induced neurogenic and inflammatory factors. Neural control <strong>of</strong> airway<br />
caliber is far from being simple and it is likely to contribute to airway<br />
narrowing and bronchial hyper-responsiveness. Myelinated and<br />
nonmyelinated nerve fibers (C fibers) are involved in the sensory irritation<br />
response and their stimulation may result in release <strong>of</strong> specific<br />
neuropeptides, known to be potent releasers <strong>of</strong> mediators from airway<br />
mast cells (Barnes et al., 1991a, b; Nielsen, 1991). Specific neuropeptides<br />
are also known to attract eosinophils which can be stimulated to release<br />
cytotoxic mediators that may exacerbate these pseudoallergic-like<br />
responses even further. Experimental and clinical studies have intimated<br />
that there is reason to suspect that acute exposure to brief high-level<br />
concentrations <strong>of</strong> asthmagenic chemicals and the development <strong>of</strong> increased<br />
airway hyperresponsiveness are associated. Thus, it could be assumed that<br />
specific mast cell sensitization—in combination with neurogenic stimuli—<br />
amplify the inflammatory process and airway hyperresponsiveness. The<br />
corresponding increase in vagal activity would increase reflex release <strong>of</strong><br />
acetylcholine and, correspondingly, may enhance airway responsiveness<br />
following the exogenous administration <strong>of</strong> cholinergic agents.<br />
Animal models <strong>of</strong> airway inflammation might allow us to investigate this<br />
relationship further. Models <strong>of</strong> allergic pulmonary inflammation have been<br />
developed in various animal species (Kips et al., 1992), using different
method- ological approaches. In toxicology, the guinea-pig has been used<br />
for decades in order to evaluate the skin sensitizing properties <strong>of</strong> chemicals<br />
and proteins and has also been able to reproduce immediate-onset<br />
pulmonary hypersensitivity responses following inhalation <strong>of</strong> chemical<br />
haptens, their protein-conjugates or antigens. This animal model has<br />
therefore been used to disclose principles governing both the development<br />
<strong>of</strong> pulmonary hypersensitivity and airway hyperreactivity. Due to the<br />
guinea-pig’s abundant amount <strong>of</strong> smooth bronchial musculature, it is used<br />
as a physiologic elicitation model that reproduces bronchospasm upon<br />
challenge to specific or nonspecific stimuli. Other animal models designed<br />
to display many <strong>of</strong> the chronic features <strong>of</strong> hypersensitivity lung diseases<br />
characteristic <strong>of</strong> occupational asthma focus more on the induction <strong>of</strong><br />
airway inflammation, the basic prerequisite for airway hyperreactivity. It<br />
should be noted, however, that the induction <strong>of</strong> asthma in the rat model,<br />
for example, commonly requires more aggressive protocols and more<br />
elaborate techniques to classify responses when compared with the guineapig<br />
elicitation model (vide infra).<br />
The guinea-pig model<br />
J.PAULUHN 131<br />
To date, practically all such models have relied upon the use <strong>of</strong> the guineapig,<br />
a species known to be sensitive for agents inducing<br />
bronchoconstriction and in which respiratory function and respiratory<br />
hypersensitivity can be measured readily. In addition, guinea-pigs are easy<br />
to handle, relatively inexpensive, and produce consistent<br />
bronchoconstrictive reactions. The models have utilized various modes <strong>of</strong><br />
hapten or antigen administration and methods for detecting sensitization,<br />
It has been shown that guinea-pigs sensitized by inhalation exposure to<br />
either a free or a protein-bound chemical can be induced to exhibit changes<br />
in respiratory patterns following inhalation challenge with the same<br />
chemical in the free or in the form <strong>of</strong> its hapten-protein conjugate. In the<br />
guinea-pig no adjuvant is needed for successful lung sensitization. More<br />
recently it has been found that changes in sensitive respiratory parameters<br />
can also be provoked in dermally sensitized guinea-pigs by inhalation<br />
challenge with the free chemical or the hapten-protein conjugate (Botham et<br />
al., 1988; Pauluhn and Eben, 1991; Hayes et al., 1992). In attempting to<br />
derive an animal model that permits the identification <strong>of</strong> asthmagenic lowmolecularweight<br />
chemicals without the presence <strong>of</strong> overriding effects<br />
caused by toxic (irritant) airway inflammation the intradermal route <strong>of</strong><br />
induction appears to be preferable. This route <strong>of</strong> induction also minimizes<br />
the risk <strong>of</strong> potential confounding effects attributable to irritant-induced<br />
nonspecific reactive airways dysfunction as a result <strong>of</strong> previous inhalation<br />
exposures (Briatico-Vangosa et al., 1993).
132 PULMONARY HYPERREACTIVITY TO INDUSTRIAL POLLUTANTS<br />
For challenge exposures it appears to be advantageous to use the free<br />
chemical in slightly irritant concentrations rather than the proteinconjugate<br />
<strong>of</strong> the hapten. It is believed that the in vitro synthesis <strong>of</strong> the<br />
hapten-protein conjugate may not necessarily result in immunologically<br />
identical conjugates when compared with those produced under in vivo<br />
conditions. Also standardized procedures to synthesize and characterize<br />
hapten-protein conjugates <strong>of</strong> multifunctional, highly reactive chemicals are<br />
not yet established. On the other hand, an essential prerequisite for<br />
challenge exposures with the free chemical is the evaluation <strong>of</strong> the irritant<br />
threshold concentration <strong>of</strong> the hapten under investigation. The importance<br />
<strong>of</strong> concentration in distinguishing irritation from sensitization cannot be<br />
overstated and is one <strong>of</strong> the most critical determinants <strong>of</strong> this animal<br />
model.<br />
For volatile, irritant haptens the characteristic feature <strong>of</strong> upper<br />
respiratory tract irritation is the reflexively induced decrease in respiratory<br />
rate which is a common finding in laboratory rodents (Figure 9.1).<br />
Consistent with this approach, naive mice, rats and guinea-pigs were<br />
exposed for 45 min to slightly irritant concentrations <strong>of</strong> phenyl isocyanate<br />
(PI). As evident from Figure 9.1, the exposure to ca. 5 mg PI m −3 air<br />
provoked a decrease in respiratory rate <strong>of</strong> approximately 25–45%. The<br />
observation that remarkable differences in response patterns between mice,<br />
rats and guinea-pigs did not occur demon strate that irritant threshold<br />
concentrations obtained in mice may also be valid for guinea-pigs. Mainly<br />
for volatile chemicals attempts have been made to establish methods for the<br />
measurement and analysis <strong>of</strong> the irritant-induced changes in respiratory<br />
pattern in mice (Vijayaraghavan et al., 1993) and to understand the<br />
mechanisms <strong>of</strong> the irritant receptor stimulation (Nielsen, 1991). For<br />
volatile irritant haptens, such as PI, an unequivocal respiratory<br />
hypersensitivity response is characterized by a shallow rapid breathing<br />
pattern, i.e. a response opposite to that occurring as a result <strong>of</strong> upper<br />
respiratory tract irritation. For volatile irritant haptens this type <strong>of</strong><br />
breathing pattern, however, can only be obtained when using the proteinconjugate<br />
<strong>of</strong> the hapten.<br />
The interpretation <strong>of</strong> changes in respiratory pattern induced by irritant<br />
particulates is less predictable because <strong>of</strong> the size-dependent deposition <strong>of</strong><br />
particles within the respiratory tract. Irritant aerosols that evoke bronchial<br />
or pulmonary irritation may produce changes similar to those occurring<br />
following immediate-onset responses. Therefore, the selection <strong>of</strong> adequate<br />
haptenchallenge concentrations as well as the measurement <strong>of</strong> several<br />
breathing parameters is <strong>of</strong> primary importance. For such chemicals,<br />
currently the relative effectiveness <strong>of</strong> the acute high-concentration<br />
inhalation (single inhalation exposure <strong>of</strong> 15 min) and the high-dose<br />
intradermal route for sensitization <strong>of</strong> guinea-pigs had been investigated<br />
(Pauluhn and Eben, 1991; Pauluhn and Mohr, 1994). The airway function
J.PAULUHN 133<br />
Figure 9.1 Time-response curves for respiratory rate from mice, rats and guineapigs<br />
during single 45-min exposures to appoximately 5 mg m −3 phenyl isocyanate.<br />
Data were normalized on pre-exposure values during 15-min air exposure. Data<br />
points for each concentration are the mean <strong>of</strong> four animals and were averaged for<br />
45 s.<br />
<strong>of</strong> conscious guinea-pigs that were sensitized to and challenged with 4,4′diphenylmethane-diisocyanate<br />
(MDI) aerosol or trimellitic anhydride<br />
(TMA) dust as well as their corresponding proteinconjugates was<br />
monitored plethysmographically. The airway hyper-responsiveness to<br />
subsequently increased inhaled acetylcholine (ACh) concentrations was<br />
assessed 1 day after the hapten challenge (Pauluhn, 1994). In most<br />
instances, selected morphological features <strong>of</strong> the airways (increased<br />
number <strong>of</strong> eosinophils in the bronchial mucosa and lung associated lymph<br />
nodes) were also taken into account.<br />
Collectively, it was noticed that elicitation <strong>of</strong> respiratory hypersensitivity<br />
is concentration-dependent and that challenge concentrations should<br />
slightly exceed the threshold concentration for irritation. The evaluation <strong>of</strong><br />
eosinophils in subepithelial tissues and lung associated lymph nodes<br />
appears to provide an important independent adjunct to measurements <strong>of</strong><br />
respiratory function. The combined assessment <strong>of</strong> specific pathologic<br />
features such as eosinophilic infiltration and the evaluations <strong>of</strong> several
134 PULMONARY HYPERREACTIVITY TO INDUSTRIAL POLLUTANTS<br />
breathing parameters upon acetylcholine and hapten or conjugate challenge<br />
significantly enhance the diagnostic sensitivity <strong>of</strong> the guinea-pig model.<br />
From studies using single, brief high level aerosol or dust exposures for the<br />
induction <strong>of</strong> animals it can be concluded that previous high level exposures<br />
evoke bronchial hyperresponsiveness upon challenge at lower hapten<br />
concentrations when compared with intradermally sensitized animals.<br />
However, guinea-pigs sensitized intradermally to the volatile PI<br />
demonstrated remarkable immediate-type respiratory reactions only upon<br />
challenge with the conjugate and not with slightly irritant concentrations<br />
<strong>of</strong> the free PI. To study if phenyl isocyanate is capable <strong>of</strong> inducing a<br />
reactive airway or an asthma like syndrome, the subsequently described rat<br />
model was used.<br />
The rat model<br />
This animal model focuses on the induction <strong>of</strong> airway inflammation which<br />
comprises most <strong>of</strong> the characteristic features <strong>of</strong> asthma. It has been stated<br />
that respiratory hypersensitivity should depend on two separate factors:<br />
first, the degree <strong>of</strong> allergic airways, and second, the sensitivity to<br />
bronchoconstrictive mediators. Increasing evidence suggests that the<br />
eosinophils play a critical role in the pathogenesis <strong>of</strong> asthma and <strong>of</strong> other<br />
non-allergic hyperresponsive airway diseases. For the induction <strong>of</strong> the<br />
asthmatic state male rats were exposed for 2 consecutive weeks by<br />
inhalation (5 h day −1 , 5 days week −1 ). The target concentrations <strong>of</strong> phenyl<br />
isocyanate were chosen on the basis <strong>of</strong> a single 45-min exposure study<br />
which suggested that approximately 1 mg m −3 air is the irritant threshold<br />
concentration for ‘any’ duration <strong>of</strong> exposure. The 2 week repeated<br />
inhalation study was designed to assess the functional, bio chemical and<br />
morphological signs <strong>of</strong> phenyl isocyanate induced lung disease and their<br />
regression during an observation period <strong>of</strong> approximately 2 months.<br />
The most characteristic features <strong>of</strong> asthma comprise an increased influx<br />
<strong>of</strong> eosinophilic granulocytes into the tissue <strong>of</strong> the airways, secretory cell<br />
hyper-plasia and metaplasia, smooth muscle hypertrophy and hyperplasia,<br />
epithelial desquamation, airway hyperresponsiveness, and eventually partial<br />
occlusion <strong>of</strong> the airway lumen with mucus and cellular debris. The<br />
formation <strong>of</strong> mucus plugs is a regular feature <strong>of</strong> asthma and accounts for<br />
most <strong>of</strong> the clinical, biochemical and physiological abnormalities.<br />
Histopathological evaluation <strong>of</strong> the respiratory tract indicated a<br />
bronchiolitis obliterans and smooth muscle hypertrophy in rats exposed to<br />
approximately 7 mg m −3 air, whereas only minimal effects were found<br />
following 4 mg m −3 air. Lung function measurements revealed that some rats<br />
were hyperresponsive to an ACh-stimulus. As shown in Figure 9.2, also the<br />
increase in shunt blood (Q s/Q t) anddecrease in forced expiratory flow rates<br />
(MMEF) as well as mucus products (sialomucins), polymorphonuclear
cells, including eosinophils, in the bronchoalveolar lavage fluid (BALF)<br />
were consistent with an asthma like syndrome. As evident from Figure 9.2,<br />
the changes observed in rats exposed to 7 mg m −3 air did not fully regress<br />
during an observation period <strong>of</strong> approximately 2 months.<br />
Conclusion<br />
J.PAULUHN 135<br />
Figure 9.2 Relative comparison <strong>of</strong> sensitive diagnostic parameters in rats exposed to<br />
either 0 (air), 1, 4 and 7 mg PI m −3 air for 2 consecutive weeks (6 h day −1 , 5 days<br />
week −1 ). The measurements were performed in weeks 3 and 9. Abbreviations:<br />
MMEF: Maximal mid-expiratory flow rate, Q s/Q t: venous admixture, PMN:<br />
polymorphonuclear cells, Eos: eosinophilic granulocytes, Sialomucins: total sialic<br />
acid (after hydrolysis).<br />
Experimental evidence suggests that changes within the respiratory tract<br />
leading to the reactive airway dysfunction syndrome and/or asthma are<br />
fully consistent with an inflammatory response involving tissue <strong>of</strong> direct<br />
contact. The toxicity <strong>of</strong> irritant chemicals known to induce such illness is<br />
highly focal, and the variability <strong>of</strong> response in different regions <strong>of</strong> the<br />
respiratory tract could be a result <strong>of</strong> the actual concentration <strong>of</strong> the<br />
toxicant reaching various airway levels. Determination <strong>of</strong> immunologic<br />
etiology is particularly important for chemical allergy since all recognized<br />
low-molecular-weight chemical sensitizers are also respiratory irritants and<br />
in sufficient concentrations can cause airway constriction by<br />
nonimmunological mechanisms. As shown by studies using phenyl
136 PULMONARY HYPERREACTIVITY TO INDUSTRIAL POLLUTANTS<br />
isocyanate, damage <strong>of</strong> the airways is characterized by a steep concentrationresponse<br />
curve. Based on acute 45-min exposure <strong>of</strong> rats the threshold for<br />
respiratory tract irritation is approximately 1 mg m −3 . Exposures equaling<br />
this concentration were tolerated without exposure-related effects, whether<br />
exposure occurred singly for 45-min or repeatedly for 2 weeks. Marginal<br />
effects were observed at 4 mg m −3 , all effects, including mortality, were<br />
produced at 7 mg m −3 . This demonstrates that selection <strong>of</strong> appropriate<br />
exposure concentrations appears to be most critical in the rat model. The<br />
assessment <strong>of</strong> diagnostic sensitivity <strong>of</strong> the methods used to probe damage<br />
to the respiratory tract demonstrated that respiratory function data, blood<br />
gas measurements, and BALF analysis facilitate a meaningful interpretation<br />
<strong>of</strong> the effects observed and are important adjuncts to common inhalation<br />
toxicological studies on rats to describe quantitatively the diseased state <strong>of</strong><br />
the lung.<br />
The guinea-pig model is experimentally less demanding and therefore can<br />
suitably be used as a screening test for respiratory sensitization, as far as<br />
the limitations <strong>of</strong> this model are taken into account. Studies on guinea-pigs<br />
demonstrate that elicitation <strong>of</strong> respiratory hypersensitivity is<br />
challengeconcentration dependent and that the concentrations used should<br />
slightly exceed the threshold concentration for irritation to maximize the<br />
magnitude <strong>of</strong> the response. However, sensitization by inhalation may<br />
increase the susceptibility to irritant stimuli and thus confounds the<br />
selection <strong>of</strong> the most appropriate concentration for challenge. The<br />
combined approach <strong>of</strong> evaluating several breathing parameters, e.g.<br />
respiratory rate, flow- and volume-derived parameters, during both the<br />
hapten (free or conjugated) and the ACh challenge provides a promising<br />
method to distinguish specific and nonspecific hypersensitivity responses.<br />
Furthermore, it is critically important to assess the respiratory irritant<br />
potency <strong>of</strong> the compound under investigation. For potent irritant<br />
substances such as volatile isocyanates, challenge with the haptenprotein<br />
conjugate minimizes the likelihood to confound specific hypersensitivity<br />
responses with those evoked merely by irritation. Taking all imponderable<br />
factors into consideration, it appears that the guinea-pig intradermalinduction<br />
inhalation-challenge protocol is adequately susceptible to identify<br />
potent respiratory tract sensitizers. However, if the airway inflammation<br />
related features <strong>of</strong> asthma are the endpoints <strong>of</strong> primary interest other<br />
animal models appear to be more appropriate.<br />
References<br />
BARNES, P.J., BARANIUK, J.N. and BELVISI, M.G., 1991a, Neuropeptides in the<br />
respiratory tract (Part II). Am. Rev. Respir. Dis., 144, 1391–9.
J.PAULUHN 137<br />
BARNES, P.J., CHUNG, K.F., PAGE, C.P., 1991b, Pharmacology <strong>of</strong> asthma,<br />
Chapter 3, Inflammatory Mediators in Page, C.P. and Barnes, P.J. (Eds.), pp<br />
54–106. Handbook <strong>of</strong> Experimental Pharmacology, Berlin, Heidelberg, New<br />
York: Springer-Verlag.<br />
BOTHAM, P.A., HEXT, P.M., RATTRAY, N.J., WALSH, S.T. and WOOD-<br />
COCK, D.R., 1988, Sensitisation <strong>of</strong> guinea-pigs by inhalation exposure to lowmolecular-weight<br />
chemicals, Toxicol. Lett., 41, 159–73.<br />
BRIATICO-VANGOSA, G, BRAUN, C.J.L., COOKMAN, G., HOFMANN, T.,<br />
KIMBER, I., LOVELESS, S.E., MORROW, T., PAULUHN, J., SØRENSEN, T.<br />
and NIESSEN, H.J., 1993, Respiratory Allergy. ECETOC Monograph No. 19.<br />
BROOKS, S.M., WEISS, M.A. and BERNSTEIN, I.L., 1985, Reactive airway dysfunction<br />
syndrome (RADS). Persistent asthma syndrome after high level irritant<br />
exposure, Chest, 88, 376–84.<br />
HAYES, J.P., DANIEL, H.R., TEE, R.D., BARNES, P.J., NEWMAN-TAYLOR,<br />
A.J. and CHUNG, K.F., 1992, Bronchial hyperreactivity after inhalation <strong>of</strong><br />
trimellitic anhydride dust in guinea-pigs after intradermal sensitization to the<br />
free hapten, Am. Rev. Respir. Dis., 146, 1311–14.<br />
KAY, A.B., 1991, Asthma and inflammation, J. Allergy Clin. Immunol., 87, 893–<br />
910.<br />
KIPS, J.C., CUVELIER, C.A. and PAUWELS, R.A., 1992, Effect <strong>of</strong> acute and<br />
chronic antigen inhalation on airway morphology and responsiveness in<br />
actively sensitized rats, Am. Rev. Respir. Dis., 145, 1306–10.<br />
NIELSEN, G.D., 1991, Mechanisms <strong>of</strong> activation <strong>of</strong> the sensory irritant receptor by<br />
airborne chemicals, Crit. Rev. Toxicol, 21, 183–208.<br />
PAULUHN, J., 1994, Test methods for respiratory sensitization in use <strong>of</strong><br />
mechanistic information in risk assessment, EUROTOX Proceedings, Arch.<br />
Toxicol., suppl. 16, 77–86.<br />
PAULUHN, J. and EBEN, A., 1991, Validation <strong>of</strong> a non-invasive technique to<br />
assess immediate or delayed onset <strong>of</strong> airway hypersensitivity in guinea-pigs, J.<br />
Appl. Toxicol, 11, 423–31.<br />
PAULUHN, J. and MOHR, U., 1994, Assessment <strong>of</strong> respiratory hypersensitivity in<br />
guinea-pigs sensitized to diphenylmethane-4,4'-diisocyanate (MDI) and<br />
challenged with MDI, acetylcholine or MDI-albumin conjugate, <strong>Toxicology</strong> (in<br />
press).<br />
VIJAYARAGHAVAN, R., SCHAPER, M., THOMPSON, R., STOCK, M.F. and<br />
ALARIE, Y., 1993, Characteristic modifications <strong>of</strong> the breathing pattern <strong>of</strong><br />
mice to evaluate the effects <strong>of</strong> airborne chemicals on the respiratory tract, Arch.<br />
Toxicol, 67, 478–90.
10<br />
Mechanisms <strong>of</strong> Pulmonary Sensitization<br />
IAN KIMBER<br />
Zeneca Central <strong>Toxicology</strong> Laboratory, Macclesfield<br />
Introduction<br />
A wide range <strong>of</strong> chemicals is known to cause allergic contact dermatitis. It<br />
is apparent, however, that chemicals also have the potential to provoke other<br />
forms <strong>of</strong> allergy and <strong>of</strong> growing concern is pulmonary sensitization.<br />
Examples <strong>of</strong> chemicals identified as human respiratory allergens are listed<br />
in Table 10.1. Respiratory allergic hypersensitivity is characterized by<br />
pulmonary reactions which occur normally in only a proportion, and<br />
frequently in only a small proportion, <strong>of</strong> exposed individuals. In those who<br />
are sensitized, respiratory reactions can be provoked by atmospheric<br />
concentrations <strong>of</strong> the causative chemical allergen which were tolerated<br />
previously and which are without effect in the non-sensitized population<br />
(Newman Taylor, 1988). Almost invariably there is a latent period between<br />
the onset <strong>of</strong> exposure and the development <strong>of</strong> respiratory symptoms such<br />
as asthma and rhinitis.<br />
By definition, allergy, including sensitization <strong>of</strong> the respiratory tract,<br />
results from the stimulation <strong>of</strong> specific immune responses by the causative<br />
agent. Although it is assumed frequently that effective allergic sensitization<br />
<strong>of</strong> the respiratory tract results largely or wholly from inhalation exposure,<br />
this is not necessarily the case. Allergic reactions manifest in a particular<br />
organ commonly result from the local provocation by the inducing agent <strong>of</strong><br />
a systemically sensitized individual. There is no reason to suppose that the<br />
quality <strong>of</strong> immune response necessary for sensitization <strong>of</strong> the respiratory<br />
tract may not result from exposure to the chemical allergen at a different<br />
site. Consistent with this is evidence that occupational respiratory allergy<br />
may be caused by dermal contact with the chemical (Karol, 1986; Nemery<br />
and Lenaerts, 1993). Furthermore, it has been reported that respiratory<br />
rate changes can be provoked by inhalation exposure <strong>of</strong> guinea pigs<br />
sensitized previously by either topical or subcutaneous treatment with the<br />
same chemical (Karol et al., 1981; Rattray et al., 1994). Despite the fact<br />
that, in practice, pulmonary sensitization may not be caused exclusively by
Table 10.1 Chemicals identified as human respiratory allergens<br />
I.KIMBER 139<br />
inhalation <strong>of</strong> the chemical allergen, it is likely that this is an important<br />
route <strong>of</strong> exposure in the occupational setting.<br />
It is well established that respiratory sensitization caused by protein<br />
aeroallergens is effected by IgE antibody. This class <strong>of</strong> antibody in man is<br />
homocytotropic and is able to associate, via specific membrane receptors,<br />
with mast cells, including mast cells in the respiratory tract. Following<br />
subsequent exposure <strong>of</strong> the sensitized individual to the same allergen, mast<br />
cell-bound IgE is cross-linked and this, in turn, results in mast cell<br />
degranulation and the release <strong>of</strong> both preformed and newly-synthesized<br />
mediators which provoke acute inflammatory reactions. In the case <strong>of</strong><br />
sensitization <strong>of</strong> the respiratory tract caused by chemicals, however, an<br />
invariable association with the presence <strong>of</strong> specific IgE antibody has failed<br />
to emerge. Although IgE antibody specific for all recognized chemical<br />
respiratory allergens has been demonstrated, it is not uncommonly the case<br />
that individuals displaying symptoms <strong>of</strong> pulmonary hypersensitivity have<br />
been reported to lack demonstrable IgE. This may suggest that<br />
immunological processes independent <strong>of</strong> IgE antibody may play a decisive<br />
role in the induction <strong>of</strong> respiratory sensitization. An alternative explanation<br />
is that inappropriate or insensitive techniques have been employed for<br />
serological analysis and that IgE antibody may be associated more<br />
commonly than suspected previously with chemical respiratory allergy. In<br />
this context it is relevant that it has been found that positive skin prick<br />
tests can be provoked in patients sensitized to acid anhydrides who, on the<br />
basis <strong>of</strong> radioallergosorbent tests (RAST), were found to lack measurable<br />
levels <strong>of</strong> serum IgE antibody (Drexler et al., 1993). Despite the absence <strong>of</strong><br />
formal confirmation that there exists a universal causal relationship<br />
between specific IgE and pulmonary hypersensitivity induced by chemicals,
140 MECHANISMS OF PULMONARY SENSITIZATION<br />
it remains likely that this class <strong>of</strong> antibody is responsible, in at least the<br />
majority <strong>of</strong> cases, for the acute onset symptoms associated with respiratory<br />
allergy (Karol et al., 1994).<br />
The induction and regulation <strong>of</strong> IgE responses<br />
IgE antibody responses are subject to a variety <strong>of</strong> immunoregulatory<br />
control mechanisms. Chief among these are the stimulatory and inhibitory<br />
actions <strong>of</strong> cytokines which serve to influence the induction and duration <strong>of</strong><br />
IgE responses. It has been found in mice that interleukin 4 (IL-4) is<br />
necessary for the initiation and maintenance <strong>of</strong> IgE antibody production<br />
(Finkelman et al., 1988b). The essential role for this cytokine in IgE<br />
responses has been emphasized further by studies <strong>of</strong> mice homozygous for<br />
a mutation that inactivates the gene for IL-4. These animals lack detectable<br />
serum IgE and fail to mount IgE responses (Kuhn et al., 1991). Importantly,<br />
in mice which produce constitutively high levels <strong>of</strong> IL-4, significantly<br />
elevated concentrations <strong>of</strong> serum IgE are evident (Burstein et al., 1991). A<br />
balance to the promotional influence <strong>of</strong> IL-4 is provided by interferon<br />
(IFN- ), a cytokine which exerts an inhibitory affect on IgE responses<br />
(Finkelman et al., 1988a). The reciprocal antagonistic activity <strong>of</strong> these<br />
cytokines is not restricted to the mouse, IL-4 and IFN- have been found to<br />
regulate human IgE production (Del Prete et al., 1988; Pene et al., 1988).<br />
The cytokines which influence the integrity <strong>of</strong> IgE responses are the<br />
products <strong>of</strong> discrete subpopulations <strong>of</strong> T helper (Th) cells, lymphocytes<br />
characterized by possession <strong>of</strong> the CD4 membrane determinant. It has been<br />
found in both mouse and man that there exists a functional heterogeneity<br />
among Th cells. Two major populations, designated Th 1 and Th 2, have<br />
been described (Mosmann and C<strong>of</strong>fman, 1989; Romagnani, 1991). It is<br />
believed currently that these subsets represent the most differentiated forms<br />
<strong>of</strong> Th cells and develop from less mature precursors as the immune<br />
response evolves (Mosmann et al., 1991). The major functional distinction<br />
between Th 1 and Th 2 cells resides in the spectrum <strong>of</strong> cytokines which they<br />
elaborate (Mosmann and C<strong>of</strong>fman, 1989). The cytokine products <strong>of</strong><br />
murine Th 1 and Th 2 cells are displayed in Table 10.2.<br />
It has been reported previously that chemicals known to cause<br />
respiratory hypersensitivity in man induce in mice immune responses<br />
characteristic <strong>of</strong> Th 2 cell activation, stimulate the production <strong>of</strong> specific IgE<br />
antibody and cause an increase in the serum concentration <strong>of</strong> IgE.<br />
Conversely, chemical allergens considered not to cause respiratory<br />
sensitivity, but which are nevertheless able to induce skin sensitization,<br />
elicit instead Th 1-type responses. In the latter case, immune responses are<br />
characterized by comparatively high levels <strong>of</strong> IgG2a antibody (an isotype<br />
known to be upregulated by IFN- ) and the absence <strong>of</strong> specific IgE<br />
(Dearman and Kimber, 1991, 1992; Dearman et al., 1991, 1992a,c,d,
Table 10.2 The cytokine products <strong>of</strong> murine Th 1 and Th 2 cells<br />
From: Mosmann and C<strong>of</strong>fman (1989).<br />
I.KIMBER 141<br />
1994). The implication is that certain chemicals favour the development <strong>of</strong><br />
Th 2 cells which will then synthesize and secrete IL-4 and thereby encourage<br />
IgE antibody responses and mast cell sensitization. The converse is that<br />
other classes <strong>of</strong> chemical allergen preferentially stimulate Th1 cells and IFNproduction.<br />
Such conditions will be nonpermissive for IgE antibody<br />
production and cell-mediated immune responses, including contact<br />
sensitization, will be favoured instead. A selective stimulation by different<br />
classes <strong>of</strong> chemical sensitizers <strong>of</strong> divergent Th responses may provide an<br />
explanation at the cellular level for the observation that chemicals vary<br />
with respect to the nature <strong>of</strong> allergic reactions that they will elicit<br />
preferentially in man. The stimulation by chemical allergens <strong>of</strong><br />
differentiated Th cell responses may have implications for allergic disease<br />
other than the regulation <strong>of</strong> IgE antibody. It is known for instance that<br />
IL-3, IL-4 and IL-10, all <strong>of</strong> which are products <strong>of</strong> murine Th 2 cells<br />
(Table 10.2), act as mast cell growth factors or c<strong>of</strong>actors (Smith and<br />
Rennick, 1986; Thompson-Snipes et al., 1991). Moreover, IL-5 is a growth<br />
and differentiation factor for eosinophils (Yokota et al., 1987) and serves<br />
to regulate the accumulation <strong>of</strong> these cells at the site <strong>of</strong> allergeninduced<br />
hypersensitivity reactions in the respiratory tract (Gulbenkian et al., 1992).<br />
It has been found recently that the cytokines IL-3 and IL-4 also enhance the<br />
secretory activity <strong>of</strong> mast cells following activation (Coleman et al., 1993).<br />
Antagonistic and inhibitory influences <strong>of</strong> Th cell products may also affect<br />
the elicitation <strong>of</strong> allergic reactions. It has been found that IFN- not only<br />
suppresses the secretory function <strong>of</strong> mast cells (Holliday et al., 1994), but<br />
also antagonizes the antigen-induced infiltration <strong>of</strong> eosinophils into the<br />
respiratory tract <strong>of</strong> sensitized mice (Iwamoto et al., 1993). Contact allergic<br />
reactions may in theory be regulated by Th 2 cytokines. It has been shown<br />
that IL-4 and IL-10 act in concert to inhibit Th 1 cell function and to
142 MECHANISMS OF PULMONARY SENSITIZATION<br />
depress cell-mediated immunity (Powrie et al., 1993) and that Il–4 is able<br />
to reduce significantly the severity <strong>of</strong> contact allergic reactions in mice<br />
(Gautam et al., 1992).<br />
Taken together the available data suggest that the selective stimulation<br />
<strong>of</strong> Th cell responses and the consequent balance created between Th 1- and<br />
Th 2-derived cytokines will have an important impact on both the induction<br />
and elicitation stages <strong>of</strong> allergy. It is perhaps not surprising, therefore, that<br />
there is increasing evidence for selective Th responses in human allergic<br />
disease. Clones <strong>of</strong> T lymphocytes specific for aeroallergens such as house<br />
dust mite and grass pollen, which cause IgE-mediated respiratory allergic<br />
reactions in susceptible individuals, have been shown to elaborate Th 2<br />
cytokines, but not IFN- (Parronchi et al., 1991). A predominance <strong>of</strong> the<br />
Th 2-type cells has been found at sites <strong>of</strong> skin reactions in atopic individuals<br />
(Kay et al., 1991) and increased numbers <strong>of</strong> IL-4 + T lymphocytes have been<br />
identified in the nasal mucosa in allergen-induced rhinitis (Ying et al., 1994).<br />
By contrast, human immune responses to nickel, a common cause <strong>of</strong><br />
allergic contact dermatitis, are characterized by the selective activation <strong>of</strong><br />
Th 1-type cells. Allergen-specific T lymphocyte clones isolated from the<br />
peripheral blood <strong>of</strong> patients sensitized to nickel have been found to secrete<br />
only low or undetectable amounts <strong>of</strong> IL-4 and IL-5, but high levels <strong>of</strong> IFN-<br />
(Kapsenberg et al., 1991).<br />
Although the relative contribution <strong>of</strong> Th 1 and Th 2 cells during immune<br />
responses, and in particular the relative availability <strong>of</strong> IL-4 and IFN- , is<br />
likely to play a predominant role in the regulation <strong>of</strong> IgE antibody, other<br />
factors may be relevant. Not least, the priming <strong>of</strong> Th 1 cells for the<br />
production <strong>of</strong> IFN- may in turn be dependent upon the action <strong>of</strong> another<br />
cytokine, interleukin 12 (IL-12) (Manetti et al., 1994; Morris et al., 1994;<br />
Schmitt et al., 1994). It has been demonstrated also that CD8 + T<br />
lymphocytes exert an important immunoregulatory influence on IgE<br />
responses (Kemeny et al., 1994; Renz et al., 1994), possibly via downregulation<br />
<strong>of</strong> CD4 + Th 2 cell development (Noble et al., 1993).<br />
It is clear that conditions outwith the immune system also influence the<br />
magnitude <strong>of</strong> IgE responses. Certainly genetic predisposition is an<br />
important, although poorly understood factor. In addition, there have been<br />
suggestions that cigarette smoking and exposure to certain environmental<br />
pollutants may result in increased IgE levels and may also serve to<br />
aggravate asthma (Zetterstrom et al., 1981; Muranka et al., 1986;<br />
Wardlaw, 1993).<br />
Cell-mediated immune responses in chemical respiratory<br />
allergy<br />
The elicitation <strong>of</strong> chemical respiratory hypersensitivity may be associated<br />
with both immediate-onset and late phase reactions. While IgE antibody
and local degranulation <strong>of</strong> mast cells may be necessary for acute<br />
symptoms, late asthmatic responses appearing some hours following<br />
exposure are characterized by an infiltration <strong>of</strong> mononuclear cells and<br />
increased numbers <strong>of</strong> leucocytes in bronchoalveolar lavage fluid. Chronic<br />
inflammation is an important component <strong>of</strong> asthma and, in addition to<br />
mononuclear cell accumulation, is characterized by mucus production, the<br />
destruction and sloughing <strong>of</strong> airway epithelial cells and subepithelial<br />
fibrosis secondary to collagen deposition. Eosinophils, acting together with<br />
infiltrating T lymphocytes, play a pivotal role in chronic bronchial<br />
inflammation (Corrigan and Kay, 1992). It is apparent also that the<br />
generation <strong>of</strong> eosinophilia in the respiratory tract is influenced markedly by<br />
Th cell products. As described previously, IL-5 effects the accumulation <strong>of</strong><br />
eosinophils at the site <strong>of</strong> hypersensitivity reactions in respiratory tissues,<br />
while IFN- , secondary to an inhibition <strong>of</strong> CD4 + cell infiltration,<br />
antagonizes this process (Gulbenkian et al., 1992; Iwamoto et al., 1993). It<br />
may prove that the cell-mediated immune processes relevant to the<br />
development <strong>of</strong> respiratory hypersensitivity and asthma are also a function<br />
<strong>of</strong> Th cell heterogeneity. Certainly the stimulation <strong>of</strong> Th 2 cell activation<br />
will have pr<strong>of</strong>ound effects on all stages <strong>of</strong> respiratory allergy. The<br />
infiltration <strong>of</strong> such cells into sites <strong>of</strong> encounter with inducing allergen, a<br />
process perhaps facilitated by vasodilation resulting from mast cell<br />
degranulation, will provide a local source <strong>of</strong> cytokines such as IL-4 and<br />
IL-5. Mast cell secretory activity will be potentiated by the former and<br />
eosinophil accumulation triggered by the latter. That Th 2 cells do in fact<br />
accumulate in the area <strong>of</strong> immediate-type hypersensitivity reactions is<br />
supported by the studies <strong>of</strong> Kay et al. (1991) who demonstrated that the<br />
cells infiltrating lesional skin at the sites <strong>of</strong> late phase cutaneous reactions<br />
in atopic patients produce IL-3, IL-4, IL-5 and GM-CSF, but not IFN- .<br />
Practical applications<br />
I.KIMBER 143<br />
In the course <strong>of</strong> investigations designed to examine the characteristics <strong>of</strong><br />
immune responses induced in mice by chemical sensitizers it was found<br />
that only those materials known to cause respiratory hypersensitivity in<br />
man provoked in mice a substantial increase in the serum concentration <strong>of</strong><br />
IgE; a phenomenon thought to reflect the selective stimulation <strong>of</strong> Th 2 celltype<br />
responses by this class <strong>of</strong> allergen. It was observed also that contact<br />
allergens known or suspected not to cause occupational respiratory<br />
hypersensitivity failed to result in similar changes in serum IgE levels<br />
(Dearman and Kimber, 1991, 1992; Dearman et al., 1992a,d). The<br />
differential ability <strong>of</strong> chemical respiratory and contact allergens to<br />
stimulate changes in the concentration <strong>of</strong> serum IgE in mice forms the basis<br />
<strong>of</strong> a novel approach to the identification <strong>of</strong> chemicals which have the<br />
potential to cause sensitization <strong>of</strong> the respiratory tract. This method, the
144 MECHANISMS OF PULMONARY SENSITIZATION<br />
mouse IgE test (Dearman et al., 1992b, Kimber and Dearman, 1993) is<br />
being evaluated currently in the context <strong>of</strong> internal and inter-laboratory<br />
validation studies.<br />
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KAY, A.B., YING, S., VARNEY, V., GAGA, M., DURHAM, S.R., MOQBEL, R.,<br />
WARDLAW, A.J. and HAMID, Q., 1991, Messenger RNA expression <strong>of</strong> the<br />
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KIMBER, I. and DEARMAN, R.J., 1993, Approaches to the identification and<br />
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(IFN- ) production during differentiation <strong>of</strong> human T helper (Th) cells and<br />
transient IFN- production in established Th 2 cell clones, Journal <strong>of</strong><br />
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I.KIMBER 147<br />
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11<br />
Occupational Asthma Induced by Chemical<br />
Agents<br />
C.A.C.PICKERING<br />
Wythenshawe Hospital Manchester<br />
Introduction<br />
Occupational asthma may be defined as variable airways narrowing<br />
causally related to exposure in the working environment to airborne dust,<br />
gases, vapours or fumes (Newman Taylor, 1980). This definition includes,<br />
therefore, both immunological and nonimmunological causes <strong>of</strong> asthma in<br />
the workplace. Immunological causes <strong>of</strong> asthma in general demonstrate a<br />
latent period between exposure and the development <strong>of</strong> symptoms. Once<br />
sensitisation has occurred airway responses may be seen at very low levels<br />
<strong>of</strong> exposure. Both high and low molecular weight agents may cause<br />
sensitisation. Irritantinduced occupational asthma characteristically follows<br />
within 24 h <strong>of</strong> a usually, single, high level exposure to an irritant substance<br />
and has been named reactive airways dysfunction syndrome (Brooks et al.,<br />
1985).<br />
The number <strong>of</strong> chemical agents causing occupational asthma is<br />
extensive. As new, highly reactive, chemicals are developed these numbers<br />
are likely to grow. Low molecular weight chemicals may act as haptens,<br />
reacting with body protein to form a complete antigen to which specific<br />
antibodies are formed.<br />
The incidence <strong>of</strong> occupational asthma in most countries is not known<br />
with any great accuracy, there are considerable variations in reporting<br />
systems between countries. Since many individuals with occupational<br />
asthma change jobs without a specific diagnosis being established, the<br />
published figures <strong>of</strong> incidence will be significant underestimates <strong>of</strong> the true<br />
incidences. In Japan (Kobayashi et al., 1973), the prevalence <strong>of</strong><br />
occupational asthma amongst adult male asthmatics is said to be about<br />
15%. In the UK a new reporting system has recently been established—<br />
Surveillance <strong>of</strong> Work-related and Occupational Respiratory Disease Project<br />
(SWORD). Newly diagnosed cases <strong>of</strong> workrelated respiratory disease are<br />
reported monthly by consultant chest and occupational physicians.<br />
Between 1989 and 1991, 631 cases <strong>of</strong> chemically induced occupational<br />
asthma were reported, <strong>of</strong> these 53% were associated with exposure to
150 OCCUPATIONAL ASTHMA INDUCED BY CHEMICAL AGENTS<br />
isocyanates. A similar system (SHIELD) has been established in the UK in<br />
the West Midlands (Gannon et al., 1993). They reported an incidence <strong>of</strong> 43<br />
new cases per million workers per year. Specific occupational incidences<br />
varied from 1833 per million paint sprayers to 8 per million clerks. Again<br />
more than half the cases <strong>of</strong> asthma were attributed to isocyanates.<br />
The initial diagnosis <strong>of</strong> occupational asthma is based on a workers<br />
history <strong>of</strong> respiratory symptoms improving on days away from work and<br />
when on holiday. At the onset <strong>of</strong> occupational asthma this pattern is<br />
usually present but continued exposure to the allergen leads to increasing<br />
airway reactivity. Their symptoms may then persist over weekends and be<br />
triggered by nonspecific factors outside the workplace such as exhaust<br />
fumes, aerosol sprays and perfumes. The difficulties <strong>of</strong> relying on the<br />
history alone in the diagnosis <strong>of</strong> occupational asthma has been well<br />
documented (Malo et al., 1991). A series <strong>of</strong> 162 hospital referrals to two<br />
expert physicians were initially categorised, on the basis <strong>of</strong> their histories,<br />
into highly probable, probable, uncertain, unlikely or absent occupational<br />
asthma. The diagnosis was then established by bronchial provocation<br />
testing and or serial measurements <strong>of</strong> lung function. The predictive value <strong>of</strong><br />
a physician’s assessment <strong>of</strong> occupational asthma being highly probable or<br />
probable was only 63%. This improved to 83% in the groups in whom<br />
occupational asthma was assessed as being unlikely or absent.<br />
The early identification <strong>of</strong> work-related symptoms and their subsequent<br />
investigation in the workplace is important. Only rarely, when very acute<br />
episodes <strong>of</strong> workplace asthma are described, should lung function<br />
measurements at work be avoided. While pre- and post-shift measurements<br />
<strong>of</strong> lung function may identify a work-related effect, late asthmatic<br />
responses occurring in the evening after leaving work are frequently seen in<br />
chemically induced forms <strong>of</strong> occupational asthma. Serial measurements <strong>of</strong><br />
lung function made every 2 h from waking to sleeping, both on working<br />
days and on days away from work, using a peak flow meter, will identify<br />
these late responders. The sensitivity <strong>of</strong> this type <strong>of</strong> investigation in<br />
establishing a diagnosis <strong>of</strong> occupational asthma is about 80% (Burge, 1982),<br />
this falls to 46% once the worker is started on specific treatment for his<br />
asthma, again emphasising the importance <strong>of</strong> early identification and<br />
investigation <strong>of</strong> work-related respiratory symptoms.<br />
Currently more than 140 low molecular weight chemicals have been<br />
reported to induce occupational asthma (Butcher and Salvaggio, 1986).<br />
The majority <strong>of</strong> these chemicals induce asthma by mechanisms which have<br />
yet to be identified. In a minority <strong>of</strong> instances specific IgE antibodies to the<br />
implicated chemical have been identified.<br />
Bronchial challenge tests with chemicals which are non-IgE dependent<br />
usually induce either an isolated late asthmatic response or a biphasic or<br />
dual asthmatic response. The IgE dependent responses induce immediate or<br />
dual asthmatic responses.
The most common chemical causes <strong>of</strong> occupational asthma include the<br />
iso- cyanates and the acid anhydrides. This chapter will examine these two<br />
groups in more detail.<br />
Isocyanates<br />
C.A.C.PICKERING 151<br />
The polyisocyanates and their oligomers are the most important cause <strong>of</strong><br />
chemically induced asthma. These organic compounds are synthesised by<br />
the reaction <strong>of</strong> amines with phosgene. There are a number <strong>of</strong> related<br />
compounds the most important <strong>of</strong> which are 2,4- and 2,6-toluene<br />
diisocyanate (TDI), methylene diphenyldiisocyanate (MDI), hexamethylene<br />
diisocyante (HDI), napthalene diisocyanate (NDI), isophorone diisocyanate<br />
(IPDI), and polyisocyanates derived from HDI and MDI.<br />
The incidence <strong>of</strong> occupational asthma due to diisocyanates varies widely.<br />
It is influenced by the type <strong>of</strong> compound and its vapour pressure. TDI and<br />
HDI are highly volatile at room temperature, whereas MDI has to be<br />
heated to above 60°C to volatilise. It is thought that approximately 5% <strong>of</strong><br />
an exposed working population will develop occupational asthma after<br />
exposure to TDI (Diem et al., 1982). Because <strong>of</strong> the known respiratory<br />
problems associated with exposure to isocyanates with high vapour<br />
pressure properties, new isocyanate compounds with low vapour pressure<br />
properties have been developed particularly for use in the paint spraying<br />
industry. Recent studies however continue to demonstrate significant levels<br />
<strong>of</strong> occupational asthma despite the use <strong>of</strong> recommended respiratory<br />
protection (Seguin et al., 1987, Welinder et al., 1988). Bronchial<br />
provocation studies with HDI- and MDI-derived polyisocyanates have<br />
confirmed their ability to cause occupational asthma. Airborne iso-cyanate<br />
prepolymers appear to be able to induce asthma to the same or greater<br />
frequency as isocyanate monomers.<br />
High exposures to isocyanate vapours, such as occur in a major<br />
industrial spillage, cause acute rhinitis, lacrymation, cough and wheezing<br />
leading to subsequent sensitisation. In some individuals this type <strong>of</strong><br />
exposure induces persistent asthma—reactive airways dysfunction<br />
syndrome (RADS). Respiratory sensitisation may occur at very low levels<br />
<strong>of</strong> exposure. Pepys et al. (1972) described a boat builder who became<br />
sensitised to TDI at exposure levels <strong>of</strong> between 0.00173 and 0.0018 ppm.<br />
Similarly White et al. (1980) reported respiratory symptoms and the<br />
development <strong>of</strong> IgE antibodies to TDI, in machinists manufacturing carseat<br />
covers exposed to levels <strong>of</strong> TDI <strong>of</strong> between 0.0003 and 0.003 ppm. It is<br />
more usual, in the author’s experience, for the sensitised individual to<br />
provide a history <strong>of</strong> short lived peak exposures to isocyanates which have<br />
clearly been above the current threshold limit value. These intermittent<br />
relatively high level exposures may be important in the sensitisation
152 OCCUPATIONAL ASTHMA INDUCED BY CHEMICAL AGENTS<br />
process. Once sensitised, a worker may have his symptoms initiated by very<br />
low exposure levels <strong>of</strong> isocyanates.<br />
Diisocyanate asthma is usually but not always associated with the<br />
presence <strong>of</strong> nonspecific bronchial hyperreactivity. The majority <strong>of</strong> workers<br />
who develop occupational asthma remain symptomatic requiring regular<br />
treatment permanently after cessation <strong>of</strong> exposure (Allard et al., 1989).<br />
The duration <strong>of</strong> exposure with symptoms before diagnosis has a major<br />
influence on recovery patterns. In a group <strong>of</strong> 43 isocyanate workers with<br />
occupational asthma, those who had fully recovered were exposed with<br />
symptoms for 1.6 years, those who had improved, 2.8 years and those who<br />
had not improved, 5.4 years (Pisati et al., 1993). The resolution or<br />
improvement in occupational asthma takes place over a 2 year period after<br />
cessation <strong>of</strong> exposure, symptoms still present at 2 years should be regarded<br />
as permanent. Most epidemiological studies have not identified any specific<br />
risk factors including atopic status, smoking or nonspecific bronchial<br />
hyperreactivity.<br />
The laboratory identification <strong>of</strong> specific antibodies to diisocyanates has<br />
proved <strong>of</strong> very limited value. Diisocyanate specific IgE is demonstrable in<br />
only 10–20% <strong>of</strong> sensitised individuals and have also been identified in<br />
individuals with no history <strong>of</strong> asthma (Butcher et al., 1983). Similarly<br />
specific IgG antibodies to diisocyanates have been described in workers<br />
both with and without evidence <strong>of</strong> disease.<br />
At the present time the recommended long-term exposure limit (8 h<br />
TWA reference period) for diisocyanates is 0.02 mg m −3 and the short-term<br />
exposure limit (10 min reference period) is 0.07 mg m −3 in the UK. There is<br />
discussion at the present time as to whether levels should be lower in order<br />
to prevent the development <strong>of</strong> diisocyanate asthma. However since most<br />
workers with diisocyanate airways disease describe exposures in excess <strong>of</strong><br />
the current recommended exposure levels the prevalence <strong>of</strong> occupational<br />
asthma in a workforce without such exposures is not known. There need to<br />
be improvements in hygiene control to prevent these peak exposures to<br />
isocyanates.<br />
Acid anhydrides<br />
The acid anhydrides are a group <strong>of</strong> low molecular weight chemicals used as<br />
curing agents in the production <strong>of</strong> epoxy and alkyd resins and in the<br />
production <strong>of</strong> plasticisers such as dioctyl phthalate. Acid anhydrides exert<br />
diverse effects on man both as sensitisers, irritants or both. The most<br />
frequently used anhydrides, all <strong>of</strong> which have been described causing<br />
occupational asthma, are phthalic anhydride (PA), trimellitic anhydride<br />
(TMA), tetrachlorophthalic anhydride (TCPA) and maleic anhydride<br />
(MA). In addition himic anhydride and pyromellitic dianhydride (PMDA)<br />
have been described as causing asthma.
The direct toxicity <strong>of</strong> anhydrides involves irritation <strong>of</strong> the mucus<br />
membranes and skin which may result in eye lesions, epistaxis, pulmonary<br />
congestion, haemoptysis and skin burns (Venables, 1989).<br />
Occupational asthma is most frequently reported due to PA, less<br />
commonly to TMA, TCPA and MA and finally there are single case reports<br />
<strong>of</strong> asthma due to HA, HHPA and PMDA (Venables, 1989). A second type<br />
<strong>of</strong> response to acid anhydrides has also been described and is termed the<br />
‘late respiratory systemic syndrome’ (LRRS). This is characterised by the<br />
development <strong>of</strong> influenzal type symptoms, fever, generalised acheing and<br />
malaise, late in the working shift or in the evening after work. These<br />
symptoms may occur in isolation or in association with asthma. It is not<br />
clear whether this response is immunologically mediated or a nonspecific<br />
response to high levels <strong>of</strong> anhydride exposure. Lastly, exposure to TMA,<br />
probably at high exposure levels, has been described as causing severe<br />
pulmonary haemorrhage requiring both blood transfusion and mechanical<br />
ventilation (Rivera et al., 1989).<br />
Serum IgE and IgG antibodies to acid anhydrides have been identified.<br />
IgE antibodies appear to be more specifically associated with occupational<br />
asthma. Howe et al. (1983) reported seven cases <strong>of</strong> TCPA asthma all <strong>of</strong><br />
whom had IgE antibody to TCPA, compared with 8% <strong>of</strong> 300 exposed<br />
workers without TCPA asthma; 29% <strong>of</strong> this exposed nonasthmatic<br />
population had IgG antibodies to TCPA.<br />
The exposure levels <strong>of</strong> acid anhydrides that initiate sensitisation are<br />
poorly understood. TMA at levels <strong>of</strong> 1.7–4.7 mg m −3 (Zeiss et al., 1977)<br />
and 0.007–2.1 mg m −3 (Bernstein et al., 1983; McGrath et al., 1984) have<br />
been described causing occupational asthma. PA at 0.03–15 mg m −3 has<br />
also been reported as causing asthma (Wernfors et al., 1986). As in other<br />
forms <strong>of</strong> occupational asthma, the early identification <strong>of</strong> cases <strong>of</strong> acid<br />
anhydride induced asthma and their removal from exposure is <strong>of</strong> prime<br />
importance.<br />
Reactive airways dysfunction syndrome<br />
C.A.C.PICKERING 153<br />
Reactive airways dysfunction syndrome (RADS) or irritant-induced asthma<br />
was first described in 1985 (Brooks et al., 1985). The criteria used in<br />
diagnosis include a high level exposure to an irritant fume, vapour or smoke,<br />
the development <strong>of</strong> respiratory symptoms within minutes or hours <strong>of</strong><br />
exposure, in an individual with no previous history <strong>of</strong> respiratory symptoms,<br />
with persistence <strong>of</strong> symptoms and physiological abnormalities for more<br />
than 1 year. A variety <strong>of</strong> different chemical exposures have been described<br />
inducing this syndrome including: chlorine (Moore and Sherman, 1991),<br />
glacial acetic acid (Kern, 1991), hydrochloric acid (Promisl<strong>of</strong>f et al., 1990)<br />
and miscellaneous chemical exposures (Brooks et al., 1985). A comparison<br />
between cases <strong>of</strong> occupational asthma and RADS (Gautrin et al., 1994)
154 OCCUPATIONAL ASTHMA INDUCED BY CHEMICAL AGENTS<br />
suggest that cases with RADS are left with less airway reversibility than<br />
occupational asthmatics. This would be consistent with the pathological<br />
findings (Boutet et al., 1993) in RADS, with more severe basement<br />
membrane thickening and bronchial wall fibrosis than is present in<br />
occupational asthma.<br />
The development <strong>of</strong> occupational asthma in any individual has<br />
potentially serious consequences both in terms <strong>of</strong> persisting disability,<br />
possible unemployment and loss <strong>of</strong> income (Gannon et al., 1993). It is<br />
incumbent on management to ensure safe working conditions with<br />
adequate control and regular monitoring <strong>of</strong> atmospheric levels <strong>of</strong> chemical<br />
agents.<br />
References<br />
ALLARD, C., CARTIER, A., GHEZZO, H. and MALO, J-L., 1989, Occupational<br />
asthma due to various agents. Absence <strong>of</strong> clinical and functional improvement<br />
at an interval <strong>of</strong> four or more years after cessation <strong>of</strong> exposure, Chest, 96,<br />
1046–9.<br />
BERNSTEIN, D.I., ROACH, D.E., MCGRATH, K.G., LARSEN, R.S., ZEISS, C. R.<br />
and PATTERSON, R., 1983, The relationship <strong>of</strong> airborne trimellitic<br />
anhydrideinduced symptoms and immune responses, J. Allergy Clin.<br />
Immunol., 72, 709–13.<br />
BOUTET, M., BOULET, L.-P., MALO, J.L., CARTIER, A., CÔTÉ, J., LEBLANC,<br />
C., MILOT, J. and LAVIOLETTE, M., 1993, Morphological evidence <strong>of</strong><br />
modified contractile properties <strong>of</strong> airways in occupational asthma and reactive<br />
airways dys-function syndrome, Am. Rev. Respir. Dis., 147, A113.<br />
BROOKS, S.M., WEISS, M.A. and BERNSTEIN, I.L., 1985, Reactive airways<br />
dysfunction syndrome. Case reports <strong>of</strong> persistent airways hyperreactivity<br />
following high-level irritant exposures, J. Occup. Med., 27, 473–6.<br />
BURGE, P.S., 1982, Single and serial measurements <strong>of</strong> lung function in the<br />
diagnosis <strong>of</strong> occupational asthma, Eur. J. Resp. Dis., 63 (suppl. 123), 47–9.<br />
BUTCHER, B.T. and SALVAGGIO, J.E., 1986, Continuing medical education—<br />
occupational asthma, J. Allergy Clin. Immunol. 78, 547–9.<br />
BUTCHER, B.T., O’NEIL, C.E., REED, M.A. and SALVAGGIO, J.E., 1983, Radioallergosorbent<br />
testing with p-tolyl monoisocyanate in toluene diisocyanate<br />
workers, Clin. Allergy., 13, 31–4.<br />
DIEM, J.E., JONES, R.N., HENDRICK, D.J., GLINDMEYER, H.W.,<br />
DHARMARAJAN, V., BUTCHER, B.T., SALVAGGIO.J.E., and WEILL, H.,<br />
1982, Five year longitudinal study <strong>of</strong> workers employed in a new toluene<br />
diisocyanate manufacturing plant, Am. Rev. Respir. Dis., 126, 420–8.<br />
GANNON, P.F.G. and BURGE, P.S., 1993, The SHIELD scheme in the West<br />
Midlands region, United Kingdom, Brit. J. Ind. Med., 50, 791–6.<br />
GANNON, P.F.G., WEIR, D.C., ROBERTSON, A.S. and BURGE, P.S., 1993,<br />
Health, employment, and flnancial outcomes in workers with occupational<br />
asthma, Brit. J. Ind. Med., 50, 491–6.
C.A.C.PICKERING 155<br />
GAUTRIN, D., BOULET, L.-P., BOUTET, M., DUGAS, M., BHÉRER, L.,<br />
L’ARCHEVÊQUE, J., LAVIOLETTE, M., CÔTÉ, J. and MALO, J.-L., 1994,<br />
Is reactive airways dysfunction syndrome a variant <strong>of</strong> occupational asthma? J.<br />
Allergy Clin. Immunol, 93, 12–22.<br />
HOWE, W., VENABLES, K.M., TOPPING, M.D., DALLY M.B., HAWKINS, R.,<br />
LAW, J.S. and NEWMAN TAYLOR, A.J., 1983, Tetrachlorophthalic<br />
anyhdride asthma: evidence for specific IgE antibody, J. Allergy Clin.<br />
Immunol., 71, 5–11.<br />
KERN, D.G., 1991. Outbreak <strong>of</strong> the reactive airways dysfunction syndrome after a<br />
spill <strong>of</strong> glacial acetic acid, Am. Rev. Respir. Dis. 144, 1058–64.<br />
KOBAYASHI, S.Y., YAMAMORA, Y., FRICK, O.L., HORIUCHI, S.<br />
KISHIMOTO, T. and MIYAMOTO, T., 1973, Occupational asthma due to the<br />
inhalation <strong>of</strong> pharmaceutical dusts and other chemical agents with some<br />
reference to other occupational asthma in Japan, Proc. VIII Int. Congr.<br />
Allergology, Tokyo, October 1973, pp. 124–32. Amsterdam: Excerpta<br />
Medica.<br />
MALO, J.L., GHEZZO, H., L’ARCHEVÊQUE, LAGIER, F. and CARTIER, A.,<br />
1991, Is the clinical history a satisfactory means for diagnosing occupational<br />
asthma? Am. Rev. Respir. Dis., 143 528–32.<br />
MCGRATH, K.G., ROACH, D., ZEISS, C.R. and PATTERSON, R., 1984,<br />
Four year evaluation <strong>of</strong> workers exposed to trimellitic anhydride: a brief<br />
report, J. Occup. Med., 26, 671–5.<br />
MOORE, B.B. and SHERMAN, M., 1991, Chronic reactive airway disease<br />
following acute chlorine gas exposure in an asymptomatic atopic patient,<br />
Chest, 100, 855–6.<br />
NEWMAN TAYLOR, A.J., 1980, Occupational asthma, Thorax, 35, 241–5.<br />
PEPYS, J., PICKERING, C.A .C., BRESLIN, A.B.X. and TERRY D.J., 1972,<br />
Asthma due to inhaled chemical agents—tolylene diisocyanate, Clin. Allergy,<br />
2. 225–36.<br />
PISATI, G., BARUFFINI, A. and ZEDDA, S., 1993, Toluene diisocyanate induced<br />
asthma: outcome according to persistence or cessation <strong>of</strong> exposure, Brit. J.<br />
Ind. Med., 50, 60–4.<br />
PROMISLOFF, R.A., PHAN, A., LENCHNER, G.S. and CICHELLI, A.V., 1990,<br />
Reactive airway dysfunction syndrome in three police <strong>of</strong>ficers following a<br />
roadside chemical spill, Chest, 98, 928–9.<br />
RIVERA, M., NICOTRA, M.B., BYRON, G.E., PATTERSON, R., YAWN, D.H.,<br />
FRANCO, M., ZEISS, C.R. and GREENBERG, S.D., 1981, Trimellitic<br />
anhydride toxicity: a cause <strong>of</strong> acute multisystem failure, Arch. Intern. Med.,<br />
141, 1071– 4.<br />
SEGUIN, P., ALLARD, A., CARTIER, A. and MALO, J.-L., 1987, Prevalence <strong>of</strong><br />
occupational asthma in spray painters exposed to several types <strong>of</strong> isocyanates,<br />
including polymethylene polyphenylisocyanate, J. Occup. Med., 29, 340–4.<br />
VENABLES, K.M., 1989, Low molecular weight chemicals, hypersensitivity, and<br />
direct toxicity: the acid anhydrides, Brit. J. Ind. Med., 46, 222–32.<br />
WELINDER, H., NIELSEN, J., BENSRYD, I. and SKERFVING, S., 1988, IgG<br />
antibodies against polyisocyanates in car painters, Clin. Allergy, 18, 85–93.
156 OCCUPATIONAL ASTHMA INDUCED BY CHEMICAL AGENTS<br />
WERNFORS, M., NIELSEN, J. and SKERFVING, S., 1986, Phthalic<br />
anhydrideinduced occupational asthma, Int. Arch. Allergy Appl. Immunol. 79,<br />
77–82.<br />
WHITE, W.G., SUGDEN, E., MORRIS, M.J. and ZAPATA, E, 1980, Isocyanateinduced<br />
asthma in a car factory, Lancet, 1, 756–60.<br />
ZEISS, C.R., PATTERSON, R., PRUZANSKY, J.J., MILLER, M.M.,<br />
ROSENBERG, M. and LEVITZ, D., 1977, Trimellitic anhydride-induced<br />
airway syndromes: clinical and immunological studies, J. Allergy Clin.<br />
Immunol., 60, 96–103.
PART FOUR<br />
Biomarkers and risk assessment <strong>of</strong><br />
industrial chemicals
12<br />
Biomarkers and Risk Assessment<br />
KARI HEMMINKI<br />
Karolinska Institute, Huddinge<br />
Introduction<br />
Many chemical carcinogens cause covalent DNA-binding products,<br />
adducts, which may induce mutations or other types <strong>of</strong> DNA damage in<br />
important growth-controlling genes or loci resulting in aberrant cellular<br />
growth and cancer (Harris, 1991; IARC 1992; Hemminki, 1993). Human<br />
exposure to compounds such as polycyclic aromatic hydrocarbons (PAH)<br />
can be determined, for example, by ambient air, biological or DNA adduct<br />
monitoring. The usefulness <strong>of</strong> a method for the determination <strong>of</strong> DNA<br />
adducts in human biomonitoring requires high sensitivity because the levels<br />
<strong>of</strong> adducts are low. Here the primary focus is on the assessment <strong>of</strong><br />
exposure using the above indicators in industries where high exposure to<br />
PAHs occur, such as iron founding, coke production, aluminium<br />
production, garage work and engine overhauling with exposure to used<br />
lubricating oils.<br />
Biomonitoring <strong>of</strong> PAH exposure<br />
Literature on the application <strong>of</strong> DNA adduct studies in humans is extensive<br />
(Beach and Gupta, 1992; IARC, 1993, 1994; Hemminki et al., 1993a;<br />
Hemminki, 1994). A large majority <strong>of</strong> the 32 P-postlabelling studies on<br />
human samples focus on tobacco smoking, occupational exposures and<br />
cancer chemotherapy patients. Most occupational exposures studied relate<br />
to complex mixtures, including polycyclic aromatic hydrocarbons (PAHs).<br />
In exposure to complex mixtures multiple radioactive spots (called diagonal<br />
radioactive zones, DRZ) are detected. The adduct spots cannot be<br />
definitively identified nor quantitated. As it has turned out that for many<br />
adducts labelling is not completed, even among structural analogues such<br />
as PAHs, the adduct levels measured are likely to be underestimates<br />
(Segerbäck and Vodicka, 1993).
Table 12.1 Exposure and aromatic adducts in occupational populations, presented in simplified tabulated form<br />
K.HEMMINKI 159<br />
Notes:<br />
a Total white blood cells.<br />
b Lymphocytes.<br />
c Not given.<br />
d No data.<br />
The types <strong>of</strong> occupational groups studied by postlabelling include<br />
foundry, coke oven and aluminium workers, ro<strong>of</strong>ers, garage and terminal<br />
workers, car mechanics and chimney sweeps. All these groups have had an
160 BIOMARKERS AND RISK ASSESSMENT<br />
increased risk <strong>of</strong> lung cancer. As, however, epidemiological studies relate to<br />
exposure a few decades earlier the risks <strong>of</strong> present exposures can only be<br />
predicted. The levels <strong>of</strong> aromatic adducts are elevated in white blood cells<br />
or lymphocytes in many <strong>of</strong> these groups. The reported total aromatic<br />
adduct levels usually range between 1 and 10 adducts per 10 8 nucleotides.<br />
Long-lived lymphocytes tend to have higher levels <strong>of</strong> adducts than shortlived<br />
granulocytes (Savela and Hemminki, 1991; Grzybowska et al., 1993).<br />
As a rule <strong>of</strong> thumb, it can be assumed that in a steady-state (i.e. long term<br />
exposure) lymphocytes contribute to the level <strong>of</strong> adducts overwhelmingly.<br />
Because they represent about 25 percent <strong>of</strong> the DNA in blood, the<br />
relationship between total white blood cell (WBC) and lymphocyte DNA<br />
adducts should be about 1:4, granulocytes only contributing to the amount<br />
<strong>of</strong> DNA denominator. Yet one has to be cautious in the comparison <strong>of</strong><br />
results between various assays even within a laboratory as the results may<br />
‘drift’ with time.<br />
The levels <strong>of</strong> white blood cell/lymphocyte aromatic adducts from<br />
workers in several industries, as measured by postlabelling, and as<br />
compared to ambient air concentrations <strong>of</strong> benzo (a) pyrene (BP) and 1hydroxypyrene<br />
levels are presented in Table 12.1. A boxplot presentation<br />
<strong>of</strong> the adduct levels <strong>of</strong> bus maintenance and truck terminal workers is<br />
shown in Figure 12.1 (Hemminki et al., 1994). The differences that were<br />
statistically significant from the controls were, in addition to the groups <strong>of</strong><br />
maintenance and terminal workers, garage workers and diesel forklift<br />
drivers.<br />
There does not seem to be a direct relationship between exposure and<br />
adduct levels. Electrode, coke and aluminium workers, exposed up to<br />
several 100 ng m −3 concentrations <strong>of</strong> BP, do not differ from the control<br />
more than foundry workers, exposed to less than 1/10 <strong>of</strong> the cited levels.<br />
The apparently higher level <strong>of</strong> adducts in the aluminium and electrode<br />
former workers (and controls) as compared to the other measurements, is<br />
due a method applied earlier with higher amounts <strong>of</strong> radioactive ATP. The<br />
later assays were carried out in small volumes but high concentrations <strong>of</strong><br />
ATP (Hemminki et al., 1993b; Szyfter et al., 1994). An increased level <strong>of</strong><br />
lymphocyte adducts has also been found in garage and truck terminal<br />
workers, with estimated exposures <strong>of</strong> about 10 ng m −3 (Hemminki et al.,<br />
1994). This would imply that the detection limit <strong>of</strong> the postlabelling<br />
method in humans exposed to PAHs lies somewhere between 1 and 10 ng<br />
m −3 BP. Whether diesel exhaust is a particularly potent inducer <strong>of</strong> adducts<br />
remains to be demonstrated. The differences between the exposed and the<br />
controls are statistically significant among foundry workers, all bus<br />
maintenance personnel and garage workers as a subgroup, all truck<br />
terminal workers and the diesel forklift drivers in particular. Coke<br />
workers differed significantly from the local controls in summer when
environmental pollution was low and the adduct levels in the controls were<br />
about 1/10 <strong>of</strong> their level in the winter (Grzybowska et al., 1993).<br />
Adducts and other endpoints<br />
K.HEMMINKI 161<br />
Figure 12.1 A boxplot <strong>of</strong> the white blood cell DNA adduct levels (per 10 8<br />
nucleotides) among bus maintenance and truck terminal workers and controls<br />
(Hemminki et al., 1994).<br />
It has become customary to include many types <strong>of</strong> endpoints to<br />
biomonitoring studies. The foundry study cited in Table 12.1 belongs to<br />
the most versatile <strong>of</strong> them. Exposure is measured by ambient air and 1hydroxypyrene<br />
monitoring (Santella et al., 1993). DNA adducts are<br />
assayed for by postlabelling and immunoassay. Plasma albumin PAH<br />
adducts are measured. Hypoxanthin guanine phosphoribosyl transferase<br />
(HPRT) and glycophorin A mutations are assayed for in lymphocytes and<br />
erythrocytes, respectively (Perera et al., 1993, 1994). Single-stand breaks in<br />
DNA and three types <strong>of</strong> cytogenetic parameters, chromosomal aberrations,<br />
sister chromatid exchanges and micronuclei, are analysed, in addition to<br />
genotyping <strong>of</strong> drug metabolising enzyme genes. Sampling <strong>of</strong> workers was<br />
repeated in four consecutive years, each at the same time <strong>of</strong> the year. As the<br />
last sampling was in the end <strong>of</strong> 1993, it will take some time before the<br />
complete data set will be available for analysis.<br />
In some published work from this data set an increase in DNA adducts<br />
and mutation frequency in the HPRT and glycophorin A genes was<br />
reported (Figure 12.2). Yet unreported results appear to show an increase
162 BIOMARKERS AND RISK ASSESSMENT<br />
Figure 12.2 Total white blood cell DNA adducts, determined by immunoassay,<br />
HPRT and glycophorin A mutations in foundry workers, exposed to various levels<br />
<strong>of</strong> BP (Perera et al., 1993).<br />
in singlestrand breaks while none <strong>of</strong> the cytogenetic parameters are<br />
elevated in the foundry workers.<br />
Adducts and metabolic genotypes<br />
The modulation <strong>of</strong> environmental carcinogenesis by host polymorphism in<br />
genes for xenobiotics metabolising enzymes is currently under extensive<br />
investigation. It was initially sparked by findings linking certain<br />
phenotypes <strong>of</strong> drug metabolism to cancer risk (Seidegård et al., 1986;<br />
Nebert, 1991). The enzymes <strong>of</strong> interest in the context <strong>of</strong> exposure to PAHs<br />
include cytochrome P450 CYP1A1 and glutathione transferase GST,<br />
involved in the activation and inactivation, respectively, <strong>of</strong> PAHs. By<br />
restriction enzyme mapping two allelic forms, ml and m2, and two other<br />
closely linked codons for isoleucine (Ile, linked to ml) and valine (Val,<br />
linked to m2) can be defined, where m2 and valine represent the rare<br />
mutant genotypes, associated with the inducibility <strong>of</strong> the enzyme activity<br />
(Hayashi et al., 1991). Polymorphism in GST1 involves the presence or the<br />
absence <strong>of</strong> the gene (Nakachi et al., 1992). The null genotype lacks the<br />
enzyme completely.<br />
Among chimney sweeps there was an association <strong>of</strong> the rare, inducible<br />
CYP1A1 genotype ml/m2 with low adduct levels in white blood cell DNA<br />
(Ichiba et al., 1994). In the same study an increased level <strong>of</strong> adducts was<br />
noted in the GST1-individuals. The level <strong>of</strong> DNA adducts appeared to be<br />
related to both the GST and CYP1A1 genotype (Figure 12.3). Analysis <strong>of</strong><br />
micronuclei in chimney sweeps resulted in no differences between<br />
individuals <strong>of</strong> either CYP1A1 ml/m2, m2/m2 or Ile/Val genotypes nor <strong>of</strong><br />
GST1 + or − genotypes (Carstensen et al., 1993). There was however a
correlation between white blood cell DNA adducts and micronuclei and it<br />
was stronger among the GST-individuals (Ichiba et al., 1994).<br />
How important is the role <strong>of</strong> metabolic phenotype or genotype as a<br />
predictor <strong>of</strong> cancer risk remains to be established. However it would seem<br />
prudent to assume some role as long as there is significant exposure to a<br />
carcinogen, metabolism <strong>of</strong> which is regulated by polymorphic genes. It<br />
would be important to note that the question can only be addressed if both<br />
<strong>of</strong> these conditions are met. In much <strong>of</strong> the published literature there are<br />
uncertainties regarding the active agents and their metabolic routes in the<br />
tissues studied. Adjustment for a metabolic phenotype or genotype, when<br />
justified, may increase the precision in the measurement.<br />
Risk assessment<br />
K.HEMMINKI 163<br />
Figure 12.3 Total white blood cell DNA adducts, measured by postlabelling,<br />
according to CYP1A1 and GST1 genotype (Ichiba et al., 1994). Controls,<br />
Sweeps.<br />
Monitoring <strong>of</strong> DNA adducts in occupational setting has mainly been<br />
applied to workers exposed to PAHs. In the case <strong>of</strong> 32 P-postlabelling<br />
increases in the level <strong>of</strong> adducts has been noted at exposures around 10 ng<br />
BP m −3 or slightly below. This is close to the detection limit that can<br />
conveniently be attained with personal monitoring or by measuring urinary<br />
1-hydroxypyrene. As the adduct measurements also reflect some aspects <strong>of</strong><br />
metabolism and DNA repair, they extend the scope <strong>of</strong> exposure
164 BIOMARKERS AND RISK ASSESSMENT<br />
measurements to host factors that may underly individual susceptibility to<br />
cancer.<br />
It has become increasingly common to try and incorporate other<br />
endpoints to DNA adduct studies. These include metabolic parameters,<br />
discussed above, protein adducts, cytogenetic parameters and point<br />
mutations. Examples include ethylene oxide exposed workers (Tates et al.,<br />
1991) and foundry workers (Perera et al., 1993, 1994; Santella et al.,<br />
1993). In both studies several parameters were elevated. The study on<br />
chimney sweeps illustrated how the intermediary endpoint may increase<br />
precision in the measurements (cf. Figure 12.3). The initial study showed<br />
no correlation between sweeping and micronuclei even though an<br />
adjustment was made for CYP1A1 and GST genotypes (Carstensen et al.,<br />
1993). There was a moderate correlation between sweeping and white<br />
blood cell DNA adducts, and adducts and micronuclei. Both <strong>of</strong> these<br />
correlations were strengthened once GST genotype was considered (Ichiba<br />
et al., 1994).<br />
Increasing circumstantial evidence associates DNA adducts within<br />
groups to an increased risk <strong>of</strong> cancer (IARC, 1992; Hemminki, 1993).<br />
Many <strong>of</strong> the adduct studies have been carried out in occupational groups<br />
which have been at a risk <strong>of</strong> cancer based on epidemiological results. These<br />
studies may be old and relate to exposures decades ago. Even new<br />
epidemiological publications on cancer cannot accurately address<br />
exposures after about 1970. Simultaneously there have been large changes<br />
in technology and industrial hygiene, undermining the quantitative and<br />
sometimes even the qualitative findings <strong>of</strong> the old epidemiological studies.<br />
This is one justification for the biomonitoring studies.<br />
Another justification is on exposures where epidemiological studies have<br />
not been conducted or have provided inadequate results, in spite <strong>of</strong><br />
suspicions raised by short-term or animal experiments. The International<br />
Agency for Research on Cancer has pointed out this as one <strong>of</strong> the criteria<br />
to be used in the evaluation <strong>of</strong> carcinogenicity <strong>of</strong> chemicals (IARC, 1992).<br />
Styrene belongs to this group <strong>of</strong> industrial exposures, where<br />
epidemiological findings are equivocal but adduct data are available on<br />
workers (Vodicka et al., 1993).<br />
Acknowledgements<br />
The research was supported by the Swedish Medical Research Council and<br />
Work Environment Fund.
References<br />
K.HEMMINKI 165<br />
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CARSTENSEN, U., ALEXANDRIE, A.-K., HÖGSTEDT,B., RANNUG, A.,<br />
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HEMMINKI, K., AUTRUP, H. and HAUGEN, A., 1993a, Environmental<br />
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HEMMINKI, K., FÖRSTI, A., LÖFGREN, M., SÄGERBÄCK, D., VACA, C. and<br />
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ICHIBA, M., HAGMAR, L., RANNUG, A.O., HÖGSTEDT, B., ALEXANDRIE, A.-<br />
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NAKACHI, K., IMAI, K., HAYASHI, S.I., WATANABE, J., KAWAJIRI, K.,<br />
HAYASHI, S.-I., WATANABE, J. and KAWAJIRI, K., 1992, High<br />
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NEBERT, D.W., 1991, Role <strong>of</strong> genetics and drug metabolism in human cancer<br />
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ØVREBø, S., HAUGEN, A., FJELDSTAD, P.E., HEMMINKI, K. and SZYFTER, K.,<br />
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ØVREBø, S., HAUGEN, A., HEMMINKI, K., SZYFTER, K., DRABLÖS, P.A. and<br />
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SAVELA, K. and HEMMINKI, K., 1993, HPRT and glycophorin A mutations<br />
in foundry worker: relationship to PAH exposure and PAH-DNA adducts,<br />
Carcinogenesis, 14, 969–73.<br />
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K., 1994, Carcinogen-DNA adducts and gene mutations in foundry workers<br />
with changing exposure to PAH, Carcinogenesis, 15, 2905–10.<br />
SANTELLA, R.M., HEMMINKI, K., TANG, D.-L., PAIK, M., OTTMAN, R.,<br />
YOUNG, T.L., SAVELA, K., VODICKOVA, L., DICKEY, C., WHYATT, R.<br />
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polycyclic aromatic hydrocarbons in the 32 P-postlabelling assay,<br />
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HEMMINKI, K., 1994, 32 P-postlabelling analysis <strong>of</strong> DNA adducts in humans:<br />
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F.J., VAN MOSSEL, H., SCHOEMAKER, H.M., OSTERMAN-GOLKAR, S.,<br />
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13<br />
Extrapolation <strong>of</strong> Toxicity Data and Assessment<br />
<strong>of</strong> Risk<br />
NORBERT FEDTKE<br />
Hüls AG, Marl<br />
Introduction<br />
Risk assessment provides a link between scientific research and risk<br />
management, or in other words, it is ‘a method for reaching public policy<br />
decisions’ (Silbergeld, 1993). Risk assessment includes the key elements<br />
hazard identification, dose-response assessment, and exposure assessment.<br />
These elements are integrated in a risk characterization step to predict<br />
adverse effects that may occur in a given population in a particular<br />
exposure situation, <strong>of</strong>ten based on the quantification <strong>of</strong> the likelihood <strong>of</strong> this<br />
occurrence. Risk management determines whether the particular exposure<br />
situation presents an acceptable or unacceptable risk and whether it is<br />
necessary to reduce the risk by reducing the exposure. Whereas risk<br />
management has to account for public health, socio-economical factors,<br />
technical feasibility, social perceptions, governmental policy and political<br />
consequences, risk assessment should be based on scientific principles.<br />
Since for the majority <strong>of</strong> industrial chemicals no or only limited human<br />
data exist, the question <strong>of</strong> how to extrapolate the data obtained from<br />
laboratory studies in experimental animals in order to predict the effects in<br />
humans has become one important aspect in risk assessment. The final<br />
goal is either to determine a level <strong>of</strong> exposure at which there is no reasonable<br />
doubt that an adverse effect will not occur in man or to define the risk<br />
associated with this exposure level. The use <strong>of</strong> mechanistic information to<br />
provide linkages between exposure, dose to tissue, and biological responses<br />
may assist in some <strong>of</strong> the steps necessary in the process <strong>of</strong> species<br />
extrapolation. Especially the use <strong>of</strong> physiologically based pharmacokinetic<br />
models (PBPK) for some aspects <strong>of</strong> risk assessment has been promoted to<br />
reduce the uncertainty associated with the current default methods. PBPK<br />
modeling is explained in general terms and a recently developed PBPK model<br />
for 2-butoxyethanol is provided as an example to illustrate the use <strong>of</strong><br />
kinetic and mechanistic data in risk assessment.
168 EXTRAPOLATION OF TOXICITY DATA AND ASSESSMENT OF RISK<br />
Legislative background in the European Union<br />
Although the process <strong>of</strong> risk assessment is not new for the individual<br />
member states <strong>of</strong> the European Union (EU) and representatives <strong>of</strong> industry<br />
and government have been assessing risk for human health and the<br />
environment for decades, EU legislation for existing 1 and new 2 substances<br />
did not formally require a systematic risk assessment up to 1992. The<br />
situation has changed with the Seventh Amendment <strong>of</strong> the Directive on the<br />
Classification, Packaging and Labelling <strong>of</strong> Dangerous Substances (EEC,<br />
1992) and the existing substances regulation (EEC, 1993a), which address<br />
risk assessment <strong>of</strong> new and existing substances, respectively.<br />
For new chemicals, the general principles <strong>of</strong> risk assessment are defined<br />
in Commission Directive 93/67/EEC (EEC, 1993b). In addition, the<br />
Directorate General XI <strong>of</strong> the European Commission has issued a series <strong>of</strong><br />
draft guidance documents for use by the competent authorities appointed<br />
by the member states. These documents provide the technical details for the<br />
risk assessment <strong>of</strong> new substances mainly by defining the testing strategies<br />
for individual toxic endpoints. For existing chemicals, a guidance<br />
document has been drafted. However, these technical guidance documents<br />
provide only little information on how to extrapolate laboratory data to<br />
humans. It has to be assumed that the extrapolation principles in the EU<br />
member states will be based on historical approaches used by authorities in<br />
other countries.<br />
Approaches to risk assessment<br />
The final goal <strong>of</strong> species extrapolation is to define a dose or dose rate<br />
which produces no adverse effects in humans. The estimation <strong>of</strong> a human<br />
no-effectlevel may include:<br />
– determination <strong>of</strong> the appropriate animal species for extrapolation to<br />
man,<br />
– determination <strong>of</strong> the most critical effect(s) and the target organ(s),<br />
– determination <strong>of</strong> the no-observed-adverse-effect level(s) (NOAEL) <strong>of</strong><br />
this effect(s), <strong>of</strong>ten followed by<br />
– extrapolation <strong>of</strong> the NOAEL(s) from subacute or subchronic to chronic<br />
exposure (time extrapolation),<br />
– extrapolation <strong>of</strong> effects observed in a high-dose region <strong>of</strong> a dose<br />
response curve to a low-dose region,<br />
1 Listed in the European Inventory <strong>of</strong> Existing Commercial Substances<br />
(EINECS).<br />
2 Not listed in EINECS.
N.FEDTKE 169<br />
– extrapolation <strong>of</strong> effects from one route <strong>of</strong> exposure to another route,<br />
and<br />
– extrapolation <strong>of</strong> effects observed in a rather homogeneous animal<br />
population to a heterogeneous human population (interspecies<br />
extrapolation), which also has to take into account the existence <strong>of</strong><br />
subgroups regarded as more sensitive as the rest <strong>of</strong> the population<br />
(intraspecies extrapolation).<br />
Essential for all procedures used in health risk assessment is the<br />
determination <strong>of</strong> the so-called critical effect. The critical effect may be<br />
defined as the adverse effect judged to be most appropriate as the basis for<br />
the risk assessment. Hence, the first step is the review <strong>of</strong> all available data<br />
on a chemical and the assessment <strong>of</strong> the adequacy <strong>of</strong> the database for the<br />
determination <strong>of</strong> the critical effect. On the basis <strong>of</strong> the critical effect,<br />
toxicants may be divided into two classes characterized by:<br />
– a threshold <strong>of</strong> response, i.e. the adverse effect on health is not expressed<br />
until the chemical, or the ultimately toxic metabolite, reaches a<br />
threshold dose or dose rate in the target tissue, or<br />
– no threshold <strong>of</strong> response, i.e. there is no threshold exposure level below<br />
which effects will not be expressed. This implies that there is some risk at<br />
any level <strong>of</strong> exposure. Examples are genotoxic carcinogens or germ cell<br />
mutagens.<br />
Based on these classes two general approaches to health risk assessment<br />
have been used.<br />
The first approach involves the use <strong>of</strong> ‘safety factors’ applied to the<br />
NOAEL or the lowest-observed-adverse-effect level (LOAEL) <strong>of</strong> a<br />
threshold effect determined in experimental animals (safety factors are<br />
recently referred to as ‘uncertainty’ or ‘assessment’ factors). The magnitude<br />
<strong>of</strong> the uncertainty factors varies between the regulatory bodies that are<br />
concerned with risk assessment, but usually they take into account the<br />
interspecies extrapolation (default factor 10) and intraspecies extrapolation<br />
(default factor 10). The magnitude <strong>of</strong> the default factors appears to be<br />
based more on the conventional use <strong>of</strong> the decimal system than on<br />
scientific reasons and have been proposed first by Lehman and Fitzhugh<br />
(1954) for the derivation <strong>of</strong> acceptable daily intakes (ADIs) for food<br />
additives. Additional uncertainty factors may be used for extrapolation to<br />
chronic exposure from subacute or subchronic exposure, adequacy <strong>of</strong> the<br />
database, extrapolation <strong>of</strong> a LOAEL to a NOAEL and severity <strong>of</strong> effects.<br />
The resulting overall uncertainty factor <strong>of</strong>ten reaches values <strong>of</strong> 1000 or<br />
higher, which is an indication <strong>of</strong> the imprecision <strong>of</strong> the derived tolerable<br />
intake. Refined extrapolation procedures using subdivisions <strong>of</strong> the default
170 EXTRAPOLATION OF TOXICITY DATA AND ASSESSMENT OF RISK<br />
factors or different default factors have recently been published (Lewis et<br />
al., 1990; Renwick, 1991, 1993).<br />
The second approach (for non-threshold effects) also relies mainly on<br />
default assumptions for dose-response extrapolation and cross-species<br />
extrapolation. Especially cancer risk assessment has been the subject <strong>of</strong><br />
much debate and there are a number <strong>of</strong> extrapolation methods reviewed<br />
recently by Park and Hawkins (1993) and Hallenbeck (1993). The default<br />
methodology in the . has been summarized by Frederick (1993). In<br />
principle, the risk assessment is based on a chronic rodent bioassay<br />
conducted at or near the maximum tolerated dose (MTD). The lifetime<br />
constant dose rates and the tumour incidence data for the individual dose<br />
groups are used to determine the dose response by fitting the data with a<br />
computer program. The linearized multistage cancer model (LMS) is <strong>of</strong>ten<br />
used to perform this step. The LMS model extrapolates the rodent tumor<br />
data observed at the MTD to a dose with a predefined risk and the 95 per<br />
cent upper bound on the dose-response curve is calculated. The interspecies<br />
extrapolation to humans is performed by a correction factor based on body<br />
weight or body surface. Subsequently, the dose is determined that<br />
corresponds to a maximum allowable calculated upper bound on risk. The<br />
resulting number does not describe the actual human risk under low-level<br />
environmental exposure, but provides an upper bound to human risk that<br />
is assumed not to be exceeded. The actual risk may be in the range between<br />
0 and the upper bound. In the process described, the dose is defined as<br />
administered dose or inhaled concentration. As a result, the lowdose<br />
extrapolation does not take into account non-linearities in tissue dosimetry<br />
and response. In addition, the interspecies extrapolation is performed using<br />
a default approach that does not account for mechanistic species<br />
differences.<br />
Use <strong>of</strong> PBPK models in risk assessment<br />
General description<br />
Physiologically based pharmacokinetic (PBPK) models have been used<br />
increasingly over the past decade to improve several aspects <strong>of</strong> the<br />
assessment <strong>of</strong> risk associated with human exposure to chemicals. Examples<br />
are PBPK models for styrene (Ramsey and Andersen, 1984; Csanády et al.,<br />
1994), dichloromethane (Andersen et al., 1987), 1,4-dioxane (Reitz et al.,<br />
1990a), chlor<strong>of</strong>orm (Reitz et al., 1990b), ethyl acrylate (Frederick et al.,<br />
1992), methanol (Horton et al., 1992) and 1,3-butadiene (Johanson and<br />
Filser, 1993). Recent reviews <strong>of</strong> the use <strong>of</strong> PBPK models in risk assessment<br />
have been published by several authors (Frederick, 1993; Travis, 1993;<br />
Wilson and Cox, 1993; Andersen and Krishnan, 1994).
N.FEDTKE 171<br />
PBPK models are based on the blood and tissue solubility <strong>of</strong> chemicals,<br />
their metabolism in various tissues and the physiology <strong>of</strong> the organism,<br />
thus incorporating the specific physiological description <strong>of</strong> animal species as<br />
well as specific physico-chemical descriptions <strong>of</strong> agents. Uptake,<br />
distribution, metabolism and excretion are described in physiologically<br />
realistic compartments (tissue groups) using computer simulation. The<br />
compartments are linked in parallel, represent the actual mammalian<br />
architecture, and include tissues such as lung and arterial blood, fatty<br />
tissue, poorly perfused tissues (muscles, skin), richly perfused tissues<br />
(brain, kidneys, heart, endocrine gland, gastro-intestinal tract), liver as the<br />
main metabolizing tissue, and mixed venous blood. The compartments are<br />
connected by arterial and venous blood flow and are characterized by a set<br />
<strong>of</strong> mass balance differential equations. The rate constants that describe the<br />
flow <strong>of</strong> material between the tissue groups and the rate <strong>of</strong> change in the<br />
chemical concentration <strong>of</strong> each compartment are proportional to blood<br />
flow, tissue solubility and compartment volumes. The basic mathematical<br />
description <strong>of</strong> a PBPK model for a volatile compound has been provided by<br />
Ramsey and Andersen (1984) and additional details may be found in<br />
appendices <strong>of</strong> manuscripts dealing with the development <strong>of</strong> PBPK models.<br />
Estimation <strong>of</strong> the constants used in PBPK models may be based on the<br />
literature in the case <strong>of</strong> physiological parameters such as ventilation rates,<br />
cardiac output, blood flow to tissues and tissue volumes. The EPA has<br />
compiled reference values for these parameters and their scaling (Arms and<br />
Travis, 1988). Chemical specific parameters such as blood and tissue<br />
solubilities may be determined from in vitro preparation (Sato and<br />
Nakajima, 1979; Gargas et al., 1989). The biochemical constants for<br />
metabolism may be derived from in vitro studies (Reitz et al. 1988;<br />
Carfagna and Kedderis, 1992; Johanson and Filser, 1993), in vivo<br />
toxicokinetic studies (Potter and Tran, 1993; Frederick et al., 1992) or in<br />
the case <strong>of</strong> volatile substances from gas uptake studies (Gargas et al., 1986,<br />
1990; Filser, 1992).<br />
Since the tissue groups have a defined biological meaning, scaling <strong>of</strong> the<br />
associated parameters between species is possible since many <strong>of</strong> the<br />
parameters used are correlated to body weight. Cardiac output, alveolar<br />
ventilation rate and V max are scaled by the 3/4 power <strong>of</strong> body weight<br />
whereas K m is assumed to be constant across species. However, the<br />
substitution <strong>of</strong> the physiological parameters with the appropriate values<br />
characteristic for the species <strong>of</strong> interest is preferred.<br />
The development <strong>of</strong> PBPK models is an iterative process involving<br />
comparison <strong>of</strong> the model simulations with experimental data and<br />
refinement <strong>of</strong> the estimates when the model fails to accurately predict the<br />
kinetic behaviour. Different exposure scenarios can be used to predict the<br />
concentrations <strong>of</strong> the parent chemical or its metabolites in the blood or the<br />
tissues, which are the target <strong>of</strong> toxic effects. The level <strong>of</strong> glutathione
172 EXTRAPOLATION OF TOXICITY DATA AND ASSESSMENT OF RISK<br />
depletion in hepatic and extrahepatic tissues (D’Souza et al., 1988;<br />
Frederick et al., 1992; Krishnan et al., 1992), kinetic interactions <strong>of</strong> parent<br />
compounds in mixed exposures (Tardif et al., 1993) or the amount <strong>of</strong><br />
adducts formed by macromolecular binding (Krishnan et al., 1992) are<br />
predictions that may also be generated by PBPK modeling. As a result <strong>of</strong><br />
the simulations, quantitative information on the internal dose <strong>of</strong> a<br />
chemical or its metabolites in the target tissue is obtained and can replace<br />
the administered dose conventionally used in risk assessment. After<br />
validation <strong>of</strong> the PBPK models in experimental animals, human PBPK<br />
models can be developed either by allometric scaling <strong>of</strong> the physiological<br />
and biochemical parameters or preferably using the actual human<br />
parameters. Following the prediction <strong>of</strong> the target tissue dosimetry in<br />
humans, the appropriate dose surrogates are related to the effect <strong>of</strong> interest<br />
and quantitative species differences are determined. This information<br />
provides the possibility to base the species extrapolation on scientific data<br />
instead <strong>of</strong> on arbitrarily assigned default factors and as a consequence the<br />
uncertainty <strong>of</strong> the extrapolation procedures applied in conventional risk<br />
assessment may be reduced.<br />
Description and use <strong>of</strong> the PBPK model for 2butoxyethanol<br />
2-Butoxyethanol (BE) is a widely produced glycol ether used as a key<br />
ingredient in water- or solvent-based coatings, industrial and consumer<br />
cleaning products, and as solvent in a variety <strong>of</strong> products. Haemolysis was<br />
identified as most sensitive indicator <strong>of</strong> BE-induced toxicity in several<br />
species <strong>of</strong> laboratory animals and has received the most attention as a<br />
critical effect for human risk assessment (ECETOC, 1985, 1994). The<br />
experimentally determined subchronic NOAEL for the rat is 25 ppm. The<br />
major metabolite <strong>of</strong> BE is 2-butoxyacetic acid (BAA) which has been<br />
identified as the metabolite responsible for the haemolysis <strong>of</strong> red blood<br />
cells in in vitro and in vivo studies (Bartnik et al., 1987; Ghanayem et al.,<br />
1987; Ghanayem, 1989). Changes in the deformability <strong>of</strong> rat erythrocytes<br />
appear to precede haemolysis upon treatment with BAA. Treatment <strong>of</strong><br />
human erythrocytes with BAA did not induce changes in deformability<br />
(Udden and Patton, 1994; Udden, 1994). The observed species differences<br />
may be due to differences in the lipid composition <strong>of</strong> erythrocyte<br />
membranes, differences in membrane proteins associated with anion<br />
transport processes, or differences in the erythrocyte cytoskeleton (Udden,<br />
1994; Udden and Patton, 1994). Humans are most likely to be exposed to<br />
BE by the dermal or inhalation routes due to the widespread use <strong>of</strong> BE in<br />
cleaning products. Assessment <strong>of</strong> the risk resulting from BE use has to<br />
account for these routes <strong>of</strong> exposure and the formation <strong>of</strong> BAA as the<br />
active metabolite. In order to assist in the risk assessment, PBPK models
N.FEDTKE 173<br />
were developed that describe the uptake, metabolism and disposition <strong>of</strong> BE<br />
and BAA (Johanson, 1986; Corley et al., 1993, 1994; Shyr et al., 1993).<br />
The model <strong>of</strong> Corley et al. (1993, 1994) is a refinement <strong>of</strong> Johanson’s<br />
model (1986) and consists <strong>of</strong> two submodels. The first submodel describes<br />
the uptake and disposition <strong>of</strong> BE and consists <strong>of</strong> the tissue compartments<br />
rapidly perfused organs, slowly perfused organs, fat, skin, muscle,<br />
gastrointestinal tract, and liver as the metabolizing tissue. The BE<br />
submodel allows uptake via the inhalation and dermal routes and in<br />
addition provides the possibility <strong>of</strong> uptake via IV infusion and the<br />
gastrointestinal tract in order to validate the model with laboratory data.<br />
The second submodel tracks the disposition <strong>of</strong> BAA in the same tissue<br />
compartments, but the kidney was removed from the rapidly perfused<br />
organs as separate tissue to allow for the excretion <strong>of</strong> BAA metabolites.<br />
The two submodels are linked together by the metabolism <strong>of</strong> BE to BAA<br />
via a saturable enzymatic pathway catalyzed by alcohol and aldehyde<br />
dehydrogenases in the liver. Competing pathways (BE conjugation and<br />
BE O-dealkylation) are lumped together and described by an additional<br />
enzymatic pathway with Michaelis-Menten kinetics. The model assumes<br />
that BAA is bound to proteins in blood and is eliminated by a saturable<br />
process in the kidneys. The rate <strong>of</strong> BAA elimination by the kidneys is<br />
described as the sum <strong>of</strong> glomerular filtration rate <strong>of</strong> BAA and the acid<br />
transport <strong>of</strong> BAA assuming that no reabsorption occurs. The biochemical<br />
constants determined experimentally in the rat were scaled to humans by<br />
(body weight) 0.7 . In the validation process, the model successfully described<br />
a wide variety <strong>of</strong> rat and human data from different laboratories using<br />
several routes <strong>of</strong> administration.<br />
BAA was predicted to be formed more rapidly in rats compared with<br />
humans, but to be eliminated slower in humans than in rats. In summary,<br />
higher maximum concentrations <strong>of</strong> BAA in blood (C max) and also higher<br />
areas under the BAA concentration-time curves (AUC) were predicted for<br />
rats than for humans, especially as the vapour concentration was<br />
increased. For the purpose <strong>of</strong> dose-response and interspecies extrapolation,<br />
BAA-C max and BAA-AUC were used as estimates <strong>of</strong> the internal dose<br />
surrogate; C max can be related directly to the in vitro haemolysis studies<br />
with BAA and is responsive to the dose-rate. The in vitro studies performed<br />
(Bartnik et al., 1987; Ghanayem et al., 1987; Ghanayem, 1989; Udden,<br />
1994; Udden and Patton, 1994) suggest that approximately 0.2 mM BAA<br />
is required to produce slight haemolysis <strong>of</strong> rat red blood cells. At about 2<br />
mM BAA nearly complete haemolysis was observed. The model predicts<br />
for nose-only exposure that these concentrations are reached in the rat at<br />
BE exposure concentrations <strong>of</strong> about 100 ppm and 800 ppm for 6 h,<br />
respectively, which is consistent with observations in vivo (Tyler, 1984;<br />
Sabourin et al., 1992).
174 EXTRAPOLATION OF TOXICITY DATA AND ASSESSMENT OF RISK<br />
For human red blood cells, the minimum BAA-concentration necessary<br />
to induce slight haemolysis is about 40 times higher compared with rats,<br />
i.e. 8 mM. The model predicts for human nose-only exposure that the C max<br />
<strong>of</strong> BAA in blood is slightly lower than the value observed in rats at a BE<br />
exposure concentration <strong>of</strong> about 100 ppm for 6 h and is only about 50 per<br />
cent <strong>of</strong> the BAA rat blood concentration at 800 ppm. In any case, the<br />
minimum toxic concentration <strong>of</strong> approximately 8 mM BAA in human<br />
blood is not achieved.<br />
The AUC has a time component which is important since haemolysis is<br />
not an instantaneous response (Udden, 1994; Udden and Patton, 1994).<br />
With respect to the AUCs for BAA, the model predicts that the values for<br />
rat and human blood are similar up to a BE exposure concentration <strong>of</strong><br />
about 500 ppm. Higher BE concentrations cause higher AUCs for BAA in<br />
rat blood than in human blood. Thus, the model predicted a BAA-AUC in<br />
man at 22 ppm BE vapour exposure that was similar to the BAA-AUC in<br />
rats achieved at 25 ppm BE vapour exposure, the established subchronic<br />
NOAEL.<br />
The simulation <strong>of</strong> the dermal BE uptake assumed that 10 per cent <strong>of</strong> the<br />
body surface <strong>of</strong> rats and humans were exposed for 6 h to BE solutions in<br />
water (5–100 per cent) and that no losses <strong>of</strong> BE occurred from the dosing<br />
solution. The simulation predicted C max blood concentrations for BAA in<br />
rats that were highest (about 3 mM) for a 40 per cent BE solution. For<br />
humans, BAA-C max was predicted to reach about 1.3 mM for the same BEconcentration.<br />
Predicted BAA-AUCs were about tw<strong>of</strong>old higher in rats<br />
compared with humans. Under these worst-case assumptions, no BE<br />
concentration is expected to achieve BAA concentrations in human blood<br />
that would cause haemolysis.<br />
ECETOC (1994) used the described PBPK model for BE and BAA<br />
disposition in combination with mechanistic data obtained by in vitro<br />
experiments to recommend an occupational exposure limit for BE:<br />
– BAA-AUC was used as the internal dose surrogate and 22 ppm BE<br />
vapour was predicted to cause a BAA-AUC in human blood similar to<br />
the BAA-AUC in rats exposed to a BE-concentration <strong>of</strong> 25 ppm<br />
(identified as the subchronic rat NOAEL). At 22 ppm BE vapour, the<br />
BAA-C max (33 µM) is predicted to be several hundredfold below the<br />
BAA concentration that causes pre-haemolytic effects in human red<br />
blood cells (8 mM).<br />
– ECETOC did not use an uncertainty factor for intraspecies<br />
extrapolation, since the in vitro studies indicated no increased sensitivity<br />
<strong>of</strong> red blood cells from individuals regarded as susceptible to haemolytic<br />
effects such as older persons, persons with hereditary spherocytosis or<br />
sickle cell disease (Udden, 1994; Udden and Patton, 1994).
– An uncertainty factor for time extrapolation (subchronic to chronic<br />
exposure) was also not applied, since the red blood cell haemolysis was<br />
regarded as a transient phenomenon observed predominantly on the<br />
first few days <strong>of</strong> exposure thus indicating that longer exposure would<br />
not have resulted in a lower rat NOAEL.<br />
– Although there is some uncertainty about the actual magnitude <strong>of</strong> the<br />
contribution <strong>of</strong> dermal uptake to the total uptake during BE vapour<br />
exposure, ECETOC concluded that even under worst-case conditions<br />
the BAA concentrations achieved are not sufficient to cause haemolysis<br />
in man and there is no need for the adjustment <strong>of</strong> the predicted human<br />
NOAEL for route.<br />
In conclusion, an occupational exposure limit <strong>of</strong> 20 ppm (8 h TWA) was<br />
recommended, also taking into account all other effects that may be<br />
associated with BE-exposure. This value is similar to the rat NOAEL <strong>of</strong> 25<br />
ppm for the most sensitive parameter, i.e. haemolysis, and was derived<br />
using scientific data instead <strong>of</strong> applying default factors to the rat NOAEL,<br />
a procedure which would have overpredicted the human risk associated<br />
with BE-exposure.<br />
Conclusion<br />
N.FEDTKE 175<br />
The use <strong>of</strong> PBPK models and mechanistic data in risk assessment tends to<br />
reduce the uncertainties in comparison with default methodologies by<br />
replacing the administered dose with the delivered dose and also tends to<br />
reveal uncertainties concealed in default methodologies (Wilson and Cox,<br />
1993). However, there are also limitations in the development <strong>of</strong> PBPK<br />
models. One limitation is that the mechanism <strong>of</strong> the toxic effect has to be<br />
known, otherwise the replacement <strong>of</strong> the external dose by internal dose<br />
surrogates is not possible. In addition, extensive validation <strong>of</strong> the model is<br />
necessary in order to replace default approaches in risk assessment. For the<br />
time being, the development <strong>of</strong> PBPK models appears to be restricted to<br />
high production chemicals where the existing data base allows<br />
identification <strong>of</strong> an accepted mechanism <strong>of</strong> toxic action and validation <strong>of</strong><br />
the model. Concern has been expressed that the use <strong>of</strong> point estimates in<br />
PBPK modelling instead <strong>of</strong> ranges <strong>of</strong> biologically plausible values leads to<br />
an increase in the uncertainty (Portier and Kaplan, 1989). However, a<br />
recent study from the Delivered Dose Work Group <strong>of</strong> the American<br />
<strong>Industrial</strong> Health Council came to the conclusion that incorporation <strong>of</strong><br />
‘pharmacokinetic information in a risk assessment,…, leads to both a more<br />
accurate estimate <strong>of</strong> risk and a better specification <strong>of</strong> the true uncertainty’<br />
(Wilson and Cox, 1993). A detailed discussion <strong>of</strong> the sources <strong>of</strong><br />
uncertainties is also provided in this reference.
176 EXTRAPOLATION OF TOXICITY DATA AND ASSESSMENT OF RISK<br />
If the data base is sufficient, PBPK models provide scientific credibility to<br />
interspecies extrapolation, extrapolation across routes <strong>of</strong> administration,<br />
extrapolation from high-dose to low-dose and intraspecies extrapolation.<br />
Recent concepts link the original tissue dose concept <strong>of</strong> PBPK models to<br />
biologically based tissue response models, thus relating the delivered dose<br />
via the mechanism <strong>of</strong> action to the toxic response and developing<br />
integrated biological models (Conolly et al., 1988; Moolgavker et al.,<br />
1988; Cohen and Ellwein, 1990; Conolly and Andersen, 1991). Such<br />
approaches enable scientists to ask the right questions and to design new<br />
mechanistic studies that will lead toward the goal <strong>of</strong> a scientifically-based<br />
risk assessment.<br />
Acknowledgement<br />
The author thanks Richard A.Corley and the Glycol Ether Panel <strong>of</strong> the<br />
Chemical Manufactures Association for providing data on 2butoxyethanol.<br />
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RENWICK, A.G., 1993, Data derived safety for the evaluation <strong>of</strong> food additives<br />
and environmental contaminants, Food Add. Contam., 10, 337–50.<br />
SABOURIN, P.J., MEDINSKY, M.A., BIRNBAUM, L.S., GRIFFITH, W.C. and<br />
HENDERSON, R.F., 1992, Effect <strong>of</strong> exposure concentration on the<br />
disposition <strong>of</strong> inhaled butoxyethanol by F-344 rats, Toxicol. Appl. Pharmacol,<br />
114, 232–8.<br />
SATO, A. and NAKAJIMA, T, 1979, Partition coefficients <strong>of</strong> some aromatic<br />
hydrocarbons and ketones in water, blood and oil, Br. J. Ind. Med., 36, 231–<br />
4.<br />
SHYR, L.J., SABOURIN, P.J., MEDINSKY, M.A., BIRNBAUM, L.S. and<br />
HENDERSON, R.F., 1993, Physiologically based modeling <strong>of</strong> 2-butoxyethanol<br />
disposition in rats following different routes <strong>of</strong> exposure, Environ. Res., 63,<br />
202–18.<br />
SILBERGELD, E.G., 1993, Risk assessment: the perspective and experience <strong>of</strong> US<br />
Environmentalists, Environ, Health Persp., 101, 100–4.<br />
TARDIF, R., LAPARÉ, S., KRISHNAN, K. and BRODEUR, J., 1993,<br />
Physiologically based modeling <strong>of</strong> the toxicokinetic interaction between<br />
toluene and m-xylene in the rat, Toxicol. Appl. Pharmacol, 120, 266–73.<br />
TRAVIS, C.C., 1993, Interspecies extrapolation <strong>of</strong> toxicological data, in Maibach,<br />
H. I. (Ed.), CRC Series in Dermatology: Clinical and Basic Science, pp. 387–<br />
410, London: CRC Press.<br />
TYLER, T.R., 1984, Acute and subchronic toxicity <strong>of</strong> ethylene glycol monobutyl<br />
ether, Environ. Health Persp., 57, 185–91.<br />
UDDEN, M.L., 1994, Hemolysis and decreased deformability <strong>of</strong> erythrocytes<br />
exposed to butoxyacetic acid, a metabolite <strong>of</strong> 2-butoxyethanol: II. Resistance<br />
in red blood cells from humans with potential susceptibility, J. Appl. Toxicol,<br />
14, 97–102.<br />
UDDEN, M.L. and PATTON, C.S., 1994, Hemolysis and decreased deformability<br />
<strong>of</strong> erythrocytes exposed to butoxyacetic acid, a metabolite <strong>of</strong> 2-butoxyethanol:<br />
I. Sensitivity in rats and resistance in normal humans, J. Appl. Toxicol, 14, 91–<br />
6.<br />
WILSON, A. and Cox, L.A., 1993, Managing Statistical Uncertainties in PBPK<br />
Modeling, Denver, CO: Cox Associates.
14<br />
Molecular Approaches to Assess Cancer Risks<br />
ALAN S.WRIGHT, J.PAUL ASTON, NICO J.VAN SITTERT<br />
and WILLIAM P.WATSON<br />
Sittingbourne Research Centre, Sittingbourne<br />
Introduction<br />
Carcinogenesis is a complex process which is not yet fully understood.<br />
Nevertheless, it is generally accepted that carcinogenesis involves the<br />
accumulation <strong>of</strong> mutations in critical genes: proto-oncogenes and/or<br />
tumour suppressor genes. These mutations transform normal cells into<br />
‘initiated’ cells possessing the full complement <strong>of</strong> genetic changes necessary<br />
for malignancy (Figure 14.1). The critical mutations may result from<br />
exposures to radiation, to genotoxic chemicals or they may arise<br />
‘spontaneously’ as a consequence <strong>of</strong> miscoding errors during the normal<br />
replication <strong>of</strong> DNA. Concomitantly, mutations will also accumulate in<br />
other genes which, although not critical for cancer per se may,<br />
nevertheless, influence cellular character thereby contributing to the<br />
multifaceted nature <strong>of</strong> cancer.<br />
The precise nature and number <strong>of</strong> critical genetic changes required for<br />
initiation have not yet been established but will probably vary from case to<br />
case. Many researchers envisage a strict temporal sequence <strong>of</strong> genetic<br />
changes in carcinogenesis. However, it is probable that the critical<br />
mutations can occur in any sequence and at any time. Indeed, it is clear<br />
that one or more <strong>of</strong> the critical mutations can occur in parental cells.<br />
Transmission (inheritance) <strong>of</strong> these mutations either through the germ line<br />
or via somatic cell division increases the susceptibility <strong>of</strong> the progeny to<br />
carcinogens. Furthermore, it is important to note that each <strong>of</strong> the critical<br />
mutations necessary for malignancy may have a different cause. This<br />
potential for multiple causation has important implications in risk<br />
assessment (vide infra).<br />
Fully initiated cells may not automatically proliferate to form tumours.<br />
One possible explanation is that the surrounding normal cells restrain<br />
the initiated or latent cancer cell by providing essential growth regulators<br />
which are no longer produced by the initiated cell. Nevertheless, partially<br />
and fully initiated cells have a replicative and/or survival advantage over<br />
normal cells. Tissue injury caused by physical trauma, chemical agents or
Figure 14.1 Schematic representation <strong>of</strong> the carcinogenic process.<br />
A.S.WRIGHT ET AL. 181<br />
viruses may have a derestraining effect thereby triggering or facilitating the<br />
replication <strong>of</strong> partially or fully initiated cells to form benign tumours or<br />
malignant tumours. Increased functional demands may also serve to<br />
promote tumour development in affected tissues.<br />
It is clear that chemicals which promote tumour development are very<br />
important determinants <strong>of</strong> carcinogenesis. Indeed, promoting agents<br />
display a marked tendency for organotropism. Promoter action is,<br />
therefore, probably the most important determinant <strong>of</strong> the site <strong>of</strong> tumour<br />
development. Yet genotoxic chemicals which initiate the carcinogenic<br />
process are perhaps viewed with even greater concern. The reasons for this<br />
high level <strong>of</strong> concern hinge mainly on evidence that the mutagenic or<br />
initiating actions <strong>of</strong> genotoxic chemicals are additive, cumulative and<br />
essentially irreversible. Furthermore, in contrast to most other classes <strong>of</strong><br />
toxic chemicals, including promoters operating via cytotoxic mechanisms,<br />
there is no theoretical reason or experimental evidence to support the view<br />
that mutagenic actions <strong>of</strong> genotoxic chemicals are thresholded. For these<br />
reasons even very low exposures <strong>of</strong> genotoxic chemicals are viewed with<br />
concern. These concerns have focused scientific and regulatory attention on<br />
a need to develop sound approaches to manage cancer risks—particularly<br />
low level risks associated with low exposures to genotoxic chemicals<br />
encountered in the occupational or environmental settings. Indeed, apart<br />
from clinical applications, high exposures to genotoxic chemicals cannot be<br />
countenanced.
182 MOLECULAR APPROACHES TO ASSESS CANCER RISKS<br />
Management <strong>of</strong> cancer risks (key requirements)<br />
The management <strong>of</strong> toxicological risks implies a capacity to control<br />
exposures within acceptable safety limits. Effective control is, therefore,<br />
dependent not only on the qualitative detection and identification <strong>of</strong><br />
hazardous chemicals but also on a capacity to determine human exposure<br />
and to evaluate the health risks. This last requirement necessitates a<br />
knowledge <strong>of</strong> potency, i.e. quantitative human dose-response relationships.<br />
In the case <strong>of</strong> genotoxic chemicals, the relevant data reside in the very low<br />
region <strong>of</strong> the dose-response curve.<br />
The concept <strong>of</strong> acceptable risk is readily accepted when applied to the<br />
many classes <strong>of</strong> toxic chemicals operating by a thresholded mechanism<br />
indicative <strong>of</strong> the virtual absence <strong>of</strong> risk at sub-threshold doses. Absolute<br />
safety margins for genotoxic chemicals cannot be guaranteed (vide supra)<br />
leading to the adoption <strong>of</strong> conservative safety measures. Indeed, cursory<br />
analysis might suggest that quantitative risk data are not required for the<br />
effective management <strong>of</strong> cancer risks associated with genotoxic chemicals.<br />
Thus, it is generally accepted that human contact with carcinogens should<br />
be minimised. Purely qualitative identification <strong>of</strong> the hazard would permit<br />
the design <strong>of</strong> measures to limit human exposure and minimise carcinogenic<br />
impact. Indeed, a purely qualitative indication <strong>of</strong> genotoxicity can be an<br />
absolute deterrent to the development <strong>of</strong> new products. Nevertheless,<br />
certain exposures, e.g. to indigenous genotoxic chemicals and natural food<br />
components, are unavoidable. Furthermore, measures to reduce exposures<br />
to ‘avoidable’ genotoxic hazards, e.g. certain combustion products, key<br />
industrial base chemicals and intermediates, are <strong>of</strong>ten difficult and costly.<br />
Quantitative risk assessment is needed to prioritise these hazards and, most<br />
importantly, to determine safety margins. Certainly a failure to determine<br />
carcinogenic potency would lead to uncertainty about the adequacy <strong>of</strong><br />
safety margins and, probably, to unnecessary measures to further reduce<br />
exposures. Thus, despite their additive and cumulative actions, even<br />
genotoxic chemicals can pose a negligible health risk. Of course, the<br />
definition <strong>of</strong> a negligible, i.e. acceptable, risk is a socio-political judgement<br />
which nevertheless has to be realistic in the case <strong>of</strong> unavoidable hazards<br />
and achievable in the case <strong>of</strong> avoidable hazards.<br />
Detection <strong>of</strong> genotoxic carcinogens<br />
Concerns about genotoxic hazards have provided an incentive for the<br />
development <strong>of</strong> a broad range <strong>of</strong> rapid tests to detect intrinsic genotoxic<br />
activity or potential. The principal aim <strong>of</strong> these approaches is to predict<br />
carcinogenic activity or, more accurately, cancer initiating activity. The<br />
most widely used tests are the coupled microsomal-microbial mutation<br />
assays developed by Ames et al. (1973). However, such approaches are
A.S.WRIGHT ET AL. 183<br />
viewed as too remote to be <strong>of</strong> value in estimating cancer risks. The trend is<br />
towards increasingly sensitive and precise technology—particularly generic<br />
methods with potential for direct application in humans.<br />
Advances in molecular biology have permitted the development <strong>of</strong> a new<br />
generation <strong>of</strong> point mutation assays based on DNA base mismatch<br />
technology (Thilly, 1991; Lu and Hsu, 1992). This technology has a<br />
precision far exceeding that <strong>of</strong> conventional biological methods and a<br />
sensitivity permitting direct applications in humans. The full potential <strong>of</strong> this<br />
technology has not yet been realised. However, it seems probable that<br />
detection levels will ultimately obviate a need for prior phenotypic<br />
selection: paving the way to universal application. Avoidance <strong>of</strong> phenotypic<br />
selection would represent a powerful advantage over existing<br />
methodologies by providing a much more direct and reliable route to<br />
determining overall background mutation rates and increments due to<br />
specific exposures <strong>of</strong> key relevance to cancer risk assessment (vide infra).<br />
The most prospective <strong>of</strong> the current assays are those designed to detect<br />
primary DNA damage. Among these procedures, 32 P-post-radiolabelling<br />
technology developed by Randerath et al. (1981) to detect DNA adducts is<br />
by far the most sensitive. The justification for application <strong>of</strong> such a<br />
prospective approach to detect exposure to genotoxic carcinogens hinges<br />
on the causal relationship established between genotoxic activity and<br />
cancer. In general, genotoxic character is conferred by possession <strong>of</strong> a<br />
centre(s) <strong>of</strong> electrophilic reactivity. This reactivity permits the chemical to<br />
undergo chemical reactions with nucleophilic centres in the target molecule<br />
(DNA). In many instances the electrophilic centre(s) is introduced into an<br />
inactive precursor chemical by metabolic activation. Primary products, e.g.<br />
DNA adducts, formed when genotoxic chemicals react with DNA are<br />
generally promutagenic (or lethal) and their occurrence leads to an<br />
increased risk <strong>of</strong> mutation and cancer. There is no known category <strong>of</strong><br />
chemical which forms DNA adducts that can be excluded from this<br />
generalisation. Not all DNA adducts are strongly promutagenic. However,<br />
because electrophiles do not display absolute specificity in their reactions<br />
with nucleophiles, the detection <strong>of</strong> even a weakly promutagenic adduct,<br />
e.g. N 7 -alkyldeoxyguanosine, signals the formation <strong>of</strong> a more strongly<br />
promutagenic adduct, e.g. O 6 -alkyldeoxyguanosine. If follows that the<br />
detection <strong>of</strong> DNA adducts provides qualitative evidence <strong>of</strong> (human)<br />
exposure to a genotoxic carcinogen.
184 MOLECULAR APPROACHES TO ASSESS CANCER RISKS<br />
Identification <strong>of</strong> human carcinogens<br />
Classical epidemiological approaches<br />
Until very recently, epidemiological approaches to detect and identify<br />
environmental carcinogens were based exclusively on the analysis <strong>of</strong><br />
tumour incidence and chromosome aberrations in human populations.<br />
However, the endpoints <strong>of</strong> these biological methods lack the intrinsic<br />
resolving power needed to dis criminate between different contributory<br />
factors. Indeed, it is only in instances <strong>of</strong> specific, high and, <strong>of</strong>ten, localised<br />
exposures that these methods have been effective in identifying specific<br />
causative agents. Nevertheless, the results <strong>of</strong> epidemiological studies<br />
indicate that chemicals, which may include both natural and xenobiotic<br />
compounds in food, drink or in the local or general environment, play a<br />
major and broad role in the aetiology <strong>of</strong> human cancer. The identification<br />
<strong>of</strong> these chemical factors is a major goal in cancer prevention.<br />
In vitro genotoxicity assays<br />
In addition to applications in screening prospective chemical products, in<br />
vitro genotoxicity assays, particularly the Ames test, provided the first<br />
practicable, systematic approach to identify environmental carcinogens.<br />
However, this approach places very heavy demands on the time and effort<br />
required to fractionate environmental samples and test individual<br />
compounds. More importantly, however, like the animal cancer studies<br />
these assays complement or have largely supplanted, the approach is not<br />
specifically targeted towards identifying and prioritising human hazards.<br />
For example, these short-term in vitro test do not provide direct evidence<br />
<strong>of</strong> human exposure or effects.<br />
Molecular epidemiology<br />
32 P-Post-radiolabelling technology for the analysis <strong>of</strong> DNA adducts<br />
provides the basis <strong>of</strong> a very sensitive and generic approach to detect<br />
exposures to genotoxic carcinogens. This technology has universal<br />
application and can be applied to detect DNA adducts formed in<br />
laboratory species or humans during exposures to both known and, as yet,<br />
unidentified genotoxic chemicals at the low concentrations encountered in<br />
the environment and the workplace. Elucidation <strong>of</strong> the chemical structures<br />
<strong>of</strong> adducts in human DNA would provide a basis for identifying the<br />
causative agents and their sources or origins. This possibility <strong>of</strong> identifying<br />
the chemical initiators <strong>of</strong> human cancer is an exciting prospect.<br />
Unfortunately, however, these adducts are present at very low abundances<br />
and this is a major obstacle to identification. Thus, the methods for
detecting DNA adducts are much more sensitive than the physicochemical<br />
methods needed for structural characterisation. A number <strong>of</strong> strategies<br />
have been adopted in attempts to solve this problem.<br />
Protein adducts<br />
Genotoxic chemicals that react with DNA also react with nucleophilic<br />
centres in proteins and may also undergo ‘spontaneous’ and enzymecatalysed<br />
reactions with glutathione leading to the excretion <strong>of</strong> the<br />
corresponding mercapturic acids. Ins<strong>of</strong>ar as the formation <strong>of</strong> protein<br />
adducts and mercapturic acids reflect the formation <strong>of</strong> the corresponding<br />
DNA adducts, their detection may also furnish evidence <strong>of</strong> exposure to a<br />
genotoxic carcinogen.<br />
The potential for reaction <strong>of</strong> genotoxic chemicals with proteins (and<br />
glutathione) is much greater than with DNA. Furthermore, human<br />
proteins, e.g. haemoglobin, are available in much larger quantities and are<br />
more accessible than human tissue DNA. These advantages have been<br />
exploited, particularly in the pioneering work <strong>of</strong> Ehrenberg’s group, to<br />
develop a range <strong>of</strong> procedures for the qualitative and quantitative analysis<br />
<strong>of</strong> protein adducts (Osterman-Golkar et al., 1976; Calleman et al., 1978;<br />
Ehrenberg and Osterman-Golkar, 1980). (For a review <strong>of</strong> the available<br />
methods see Skipper and Naylor, 1991.) The most powerful and generic<br />
approach is undoubtedly that developed by Törnqvist et al. (1986a). An<br />
initial purification or enrichment step is key to any successful method for<br />
the analysis <strong>of</strong> low levels <strong>of</strong> organic residues. The amino-groups <strong>of</strong> the Nterminal<br />
valine residues <strong>of</strong> the α-and<br />
β-chains <strong>of</strong> human haemoglobin are<br />
major targets for reaction with a broad range <strong>of</strong> genotoxic chemicals.<br />
Törnqvist achieved selective enrichment <strong>of</strong> adducted N-terminal valine<br />
residues <strong>of</strong> haemoglobin by devising a modified Edman degradation which<br />
resulted in the scission <strong>of</strong> adducted residues whilst leaving the nonadducted<br />
N-terminal valines intact. This procedure provides the basis for<br />
identifying the adducting moieties and their quantitation by GC/MS.<br />
Applications <strong>of</strong> this technology have furnished evidence <strong>of</strong> background<br />
exposures to a range <strong>of</strong> alkylating species. Protein adduct technology has<br />
the potential for considerable further refinement. The possibility <strong>of</strong> using<br />
immunoaffinity technology to enrich both known and unidentified protein<br />
adducts is currently being explored.<br />
DNA adducts and immunoenrichment<br />
A.S.WRIGHT ET AL. 185<br />
The need for effective enrichment technology for DNA adducts is even<br />
more pressing than for protein adducts. Ideally, the enrichment procedure<br />
should be applied at the earliest possible stage <strong>of</strong> analysis. The procedure
186 MOLECULAR APPROACHES TO ASSESS CANCER RISKS<br />
should be rapid and mild in order to minimise the formation <strong>of</strong> artefacts.<br />
Currently, immunoaffinity technology holds the greatest promise.<br />
Immunoenrichment <strong>of</strong> DNA adducts necessitates antibodies possessing<br />
the appropriate specificities and affinities to permit selective binding <strong>of</strong><br />
adducts at the very low abundances encountered in hydrolysates or enzyme<br />
digests <strong>of</strong> human DNA. The immune system does not normally respond to<br />
small molecules per se. However, the system can be induced to produce<br />
effective antibodies by immunising with the small molecule (hapten)<br />
coupled to a protein. Such treatment induces a spectrum <strong>of</strong> antibodyproducing<br />
cells, each producing a specific antibody. Most <strong>of</strong> these<br />
antibodies recognise various regions (epitopes) <strong>of</strong> the carrier protein while<br />
a few may specifically recognise and bind the small molecule <strong>of</strong> interest.<br />
Suitable antibody-producing cells can be selected and cloned to provide a<br />
permanent source <strong>of</strong> homogenous antibody (monoclonal antibody, Mab).<br />
Mabs can be raised against virtually any organic chemical although some<br />
lower molecular weight compounds (
adducts are not viewed as particularly promutagenic. Nevertheless the N 7 -<br />
atom <strong>of</strong> dG residues in DNA is a major target for adduction and the<br />
detection <strong>of</strong> N 7 -dG adducts signals the production <strong>of</strong> adducts at other<br />
(more critical) sites in DNA (vide supra).<br />
In certain instances, a class <strong>of</strong> adducting moieties may possess a common<br />
structural feature that can be exploited for immunoenrichment. For<br />
example, a Mab has been raised against the major DNA adduct <strong>of</strong> benzo(a)<br />
pyrene (r-7,t-8,t-9-trihydroxy-c-10-(N 2 -deoxyguanosylphosphate)-7,8,9,<br />
10-tetrahydrobenzo(a)pyrene) in a collaborative study with Dr Baan’s<br />
group. This Mab recognises DNA adducts formed by a broad range <strong>of</strong><br />
polycyclic aromatic hydrocarbons (PCAs) including benzo(a)pyrene (BP),<br />
chrysene, benz(a)anthracene, 5-methylchrysene, picene and dibenz(a,h)<br />
anthracene. It seems probable that this Mab recognises the common<br />
trihydric alcohol structure produced when the reactive diol epoxides <strong>of</strong><br />
each <strong>of</strong> these polycyclic compounds reacts with nucleophilic centres in<br />
DNA or other macromolecules. The fact that the Mab does not bind the<br />
corresponding fluoranthene adduct is consistent with the spatial<br />
environment <strong>of</strong> the hydroxyl groups in fluoranthene-DNA adducts which<br />
is completely different from those generated from the other PCAs employed<br />
in this study.<br />
The performance <strong>of</strong> the Mab raised against the major BP-DNA adduct in<br />
the enrichment <strong>of</strong> PCA-DNA adducts is being evaluated using the<br />
immobilised Mab coupled to cyanogen bromide-activated Sepharose 4B.<br />
Results obtained to date demonstrate that the immobilised Mab selectively<br />
adsorbs the major BP-DNA adduct from DNA hydrolysates at abundances<br />
below 1 adduct per 10 9 nucleotide units. Results with the other PCA-DNA<br />
adducts are not yet available. However, the results obtained with the major<br />
BP-DNA adduct underlines the potential <strong>of</strong> immunoenrichment technology<br />
in the qualitative and quantitative analysis <strong>of</strong> adducts. Furthermore, such<br />
results provide an incentive to pursue the development <strong>of</strong> class-specific<br />
antibodies in order to permit or facilitate the identification <strong>of</strong> the chemical<br />
initiators <strong>of</strong> human cancer.<br />
Mercapturic acids<br />
A.S.WRIGHT ET AL. 187<br />
Qualitative analysis <strong>of</strong> mercapturic acids also provides a basis for<br />
identifying human exposures to genotoxic chemicals (vide supra). However,<br />
the available analytical procedures are complex and tend to lack specificity<br />
and sensitivity. During the last dozen years we have undertaken a number<br />
<strong>of</strong> studies aimed at developing compound- and class-specific antibodies to<br />
facilitate the analysis <strong>of</strong> mercapturic acids.<br />
Conventional approaches to generate antibodies to low molecular weight<br />
(MW) organic chemicals involves the covalent attachment <strong>of</strong> the small<br />
molecule (hapten) to a strongly antigenic protein, e.g. keyhole limpet
188 MOLECULAR APPROACHES TO ASSESS CANCER RISKS<br />
haemocyanin or bovine serum albumin, for the immunisation <strong>of</strong> mice. This<br />
strategy is usually effective in the case <strong>of</strong> strongly antigenic haptens, e.g.<br />
aromatic nitro compounds <strong>of</strong> PCA-DNA adducts. However, all <strong>of</strong> our<br />
attempts to use this approach to generate antibodies against relatively low<br />
MW and weakly antigenic mercapturic acids, e.g. S-(2-hydroxyethyl)-Nacetylcysteine,<br />
failed. Antibodies were generated but these were directed<br />
against the strongly antigenic carrier protein(s).<br />
Covalent binding to macromolecules is believed to provide the basis <strong>of</strong><br />
allergic responses, e.g. skin sensitisation reactions, to small molecules and,<br />
possibly, a basis for the induction <strong>of</strong> auto-immune responses. Thus, the<br />
binding <strong>of</strong> the small molecule transforms normal proteins into ‘foreign’<br />
proteins which trigger an immune response. Recently we have employed<br />
this principle in an attempt to direct the immune response specifically<br />
against mercapturic acid haptens by immunising mice with the haptens<br />
bound to a non-antigenic carrier protein, i.e. mouse serum albumin.<br />
Preliminary results indicate that this tactic has been successful. Overall the<br />
treatment induced fewer antibody-producing cells. However, the antibodies<br />
that were generated show high affinities and specificity toward model<br />
mercapturic acids including S-(2-hydroxyethyl) and S-phenylmercapturic<br />
acid. Studies are in progress to investigate the performance <strong>of</strong> these<br />
antibodies in an immunoenrichment mode.<br />
The preliminary results <strong>of</strong> our studies using non-antigenic protein<br />
carriers are very encouraging and have provided fresh insights which may<br />
assist in directing immune responses against the specific structural features<br />
<strong>of</strong> interest. Improvements in our ability to tailor the antibody will prove<br />
extremely valuable in optimising the properties <strong>of</strong> antibodies to meet<br />
specific needs, e.g. to enrich DNA, protein or mercapturic acid adducts for<br />
application in identifying the chemical initiators <strong>of</strong> human cancer and<br />
quantifying exposures to these agents.<br />
Cancer risk assessment<br />
Human exposure monitoring (determination <strong>of</strong> dose)<br />
The assessment <strong>of</strong> cancer risks posed by exposure to genotoxic chemicals<br />
has two components: determination <strong>of</strong> the dose and determination <strong>of</strong> the<br />
effect (increment in cancer incidence) caused by that dose. The introduction<br />
<strong>of</strong> the target dose concept by Ehrenberg in the early 1970s has provided the<br />
key to modern strategies to assess genotoxic risks (Ehrenberg, 1974, 1979;<br />
Ehrenberg et al., 1974). This new dose concept was developed to provide a<br />
measure <strong>of</strong> the critical dose, i.e. the dose <strong>of</strong> the ultimate genotoxic agent(s)<br />
penetrating to DNA. Target dose is much more relevant to risk assessment<br />
than is exposure dose. The determination <strong>of</strong> target dose automatically
compensates for individual or species differences in the operation <strong>of</strong><br />
metabolic and biokinetic factors that control the quantitative (and<br />
qualitative) relationships between the exposure and the dose <strong>of</strong> the ultimate<br />
toxicant delivered to the target. Measurements <strong>of</strong> target dose may be<br />
applied to improve the extrapolation <strong>of</strong> risk data from experimental<br />
models to humans and may also provide improved definitions <strong>of</strong> risks to<br />
individuals. The determination <strong>of</strong> target dose in humans may be viewed,<br />
therefore, as an approach towards direct risk monitoring as well as a more<br />
relevant approach to monitor human exposures to genotoxic chemicals.<br />
Determination <strong>of</strong> target dose<br />
The determination <strong>of</strong> target dose raises numerous technical and theoretical<br />
problems. Target dose can be determined by measuring primary products,<br />
e.g. DNA adducts, formed when genotoxic agents react with DNA. The<br />
kinetics <strong>of</strong> formation and decay <strong>of</strong> these adducts must also be determined<br />
(vide infra) in order to transform measurements <strong>of</strong> amounts <strong>of</strong> adducts into<br />
estimates <strong>of</strong> target dose. Human tissue DNA is not readily accessible for<br />
monitoring purposes: surrogate dose monitors are required. There are<br />
numerous possibilities including the determination <strong>of</strong> adducts in white<br />
blood cell DNA or <strong>of</strong> the corresponding adducts in the haemoglobin <strong>of</strong><br />
circulating erythrocytes.<br />
Such indirect approaches require validation. Haemoglobin is the most<br />
extensively studied surrogate, not only because <strong>of</strong> its accessibility and<br />
relative abundance but also because <strong>of</strong> the relative stability <strong>of</strong> haemoglobin<br />
adducts and the longevity <strong>of</strong> erythrocytes which permit retrospective<br />
estimates <strong>of</strong> dose received by the erythrocytes over a period <strong>of</strong> about 4<br />
months. Current evidence indicates that all electrophiles that undergo<br />
covalent reactions with DNA also react with haemoglobin. Furthermore<br />
the amounts <strong>of</strong> haemoglobin adducts are quantitatively related to the rates<br />
<strong>of</strong> formation <strong>of</strong> DNA adducts in the tissues. However the proportional<br />
relationships between the doses delivered to tissue DNA and to<br />
haemoglobin or to any other surrogate will vary from chemical to chemical<br />
and will have to be established using experimental models.<br />
Measurement <strong>of</strong> haemoglobin adducts<br />
A.S.WRIGHT ET AL. 189<br />
Genotoxic chemicals undergo covalent reactions with a variety <strong>of</strong><br />
nucleophilic centres in haemoglobin including the sulphydryl group <strong>of</strong><br />
cysteine, the N 1 and N 3 atoms <strong>of</strong> histidine and the amino groups <strong>of</strong> Nterminal<br />
valine residues. Ehrenberg’s group (Osterman-Golkar et al., 1976;<br />
Calleman et al, 1978; Törnqvist et al., 1986a) has pioneered the<br />
development <strong>of</strong> methods to detect, identify and quantify adducts formed at<br />
each <strong>of</strong> these centres. A review <strong>of</strong> these and methods for the analysis <strong>of</strong>
190 MOLECULAR APPROACHES TO ASSESS CANCER RISKS<br />
‘labile’ adducts formed, for example, during exposure to aromatic amines<br />
(Green et al., 1984; Albrecht and Neumann, 1985) is beyond the scope <strong>of</strong><br />
this paper (for reviews see Farmer, 1991; Skipper and Naylor, 1991).<br />
However, probably the most powerful and valuable approach was<br />
developed by Törnqvist et al. (1986a, b) who showed that adducts with the<br />
N-terminal valine residues <strong>of</strong> haemoglobin could be specifically enriched by<br />
scission in a modified Edman reaction followed by extraction. This<br />
enrichment procedure greatly facilitates sample analysis by GC/MS.<br />
Immunoassays are also being introduced as alternatives to physicochemical<br />
methods for the determination <strong>of</strong> protein adducts (Wraith et al., 1988).<br />
However, as in the case <strong>of</strong> DNA adducts, the biggest impact <strong>of</strong><br />
immunotechnology on the analysis <strong>of</strong> protein adducts will probably be in<br />
the immunoenrichment <strong>of</strong> low levels <strong>of</strong> adducts for analysis by physicochemical<br />
methods.<br />
Determination <strong>of</strong> biological effects<br />
Tumour incidence<br />
The determination <strong>of</strong> target dose is essential for assessing cancer risks<br />
posed by low-level exposures to genotoxic chemicals. The other requisite is<br />
know ledge <strong>of</strong> the human dose-carcinogenic response relationships in the<br />
low-dose range. The lack <strong>of</strong> intrinsic resolving power <strong>of</strong> classical<br />
epidemiological methods (vide supra) prevents effective applications to<br />
detect small carcinogenic effects associated with low exposures to any<br />
particular genotoxic chemical. Furthermore, the detection limits <strong>of</strong> animal<br />
cancer studies fall short <strong>of</strong> ‘acceptable’ risk limits by three to four orders <strong>of</strong><br />
magnitude (Wright, 1991). This poor sensitivity compels the use <strong>of</strong> high<br />
test doses in order to ensure that significant carcinogens do not go<br />
undetected. However, it is generally accepted that high doses <strong>of</strong> chemicals<br />
may induce tumours by non-specific mechanisms, e.g. via tissue injury and<br />
compensatory cell proliferation, that do not operate at low doses (Ames,<br />
1989; Wright, 1991). Many, if not all genotoxic chemicals induce cell<br />
injury at high (thresholded) doses. Clearly, extrapolation <strong>of</strong> such high dose<br />
risk data to the relevant low-dose range may, at the very least, lead to a<br />
gross overestimation <strong>of</strong> risk.<br />
Determination <strong>of</strong> mutagenic potency<br />
In considering the impact that a low-level exposure to a genotoxic<br />
chemical may have on cancer incidence, it is reasonable to suggest that the<br />
mutagenic propensity <strong>of</strong> the chemical, although <strong>of</strong> a low order, would<br />
nevertheless be the overriding risk factor. Thus, it is probable that any
A.S.WRIGHT ET AL. 191<br />
intrinsic promoter activity <strong>of</strong> the chemical arising, for example as a<br />
consequence <strong>of</strong> cell injury, would be negligible at low exposures—<br />
particularly when viewed in the context <strong>of</strong> the overall promoter pressure<br />
exerted on the populations at risk. Thus, it is unlikely that the added<br />
cancer risk would be greater than and may approximate to the increment in<br />
critical mutations caused by the specific exposure (Figure 14.1).<br />
Experimental studies indicate that all known categories <strong>of</strong> genotoxic<br />
agents ranging from methylating agents to polycyclic aromatic compounds<br />
can induce the critical mutations leading to malignancy (Figure 14.1).<br />
Furthermore, at low doses, the possibility that exposure to a particular<br />
genotoxic chemical would induce more than one <strong>of</strong> the critical mutations in<br />
any particular cell is extremely remote. Indeed it is probable that each<br />
critical mutation is induced by a different agent or mechanism, i.e.<br />
chemical, radiation or ‘spontaneous’. In this sense, each critical mutation<br />
would have equal status, i.e. no single event would be any more or any less<br />
critical than any other to the final outcome. Accordingly, the increment in<br />
cancer risk would equate with the increment in any decisive, e.g. oncogeneactivating,<br />
mutation in any critical gene. The induction <strong>of</strong> such mutations<br />
is almost certainly a direct function <strong>of</strong> overall mutagenic activity <strong>of</strong> the<br />
chemical, i.e. linked to the number <strong>of</strong> mutational events rather than the type<br />
<strong>of</strong> mutations induced by a given dose <strong>of</strong> the chemical. At low exposures,<br />
therefore, the increment in cancer incidence due to a specific genotoxic<br />
agent would approximate to the small increase in the total mutational load<br />
caused by the exposure, i.e. relative to the overall background level <strong>of</strong><br />
mutations due to all causes, multiplied by the overall cancer incidence in<br />
the population at risk. (The latter function introduces a measure <strong>of</strong> the net<br />
impact <strong>of</strong> promoter and anti-promoter pressure acting on initiated cells in<br />
the population at risk.) Of course risks may also be calculated on the basis<br />
<strong>of</strong> specific tumours and specific tissues.<br />
The determination <strong>of</strong> small increments in mutation associated with lowlevel<br />
exposures to genotoxic chemicals in human populations presents<br />
enormous technical problems not least due to the much larger and variable<br />
background <strong>of</strong> mutations due to all causes. Increasing the sensitivity <strong>of</strong><br />
mutation assays per se (vide supra) is unlikely to improve the situation.<br />
High resolving power is also needed to discriminate between effects due to<br />
different contributory factors. Nevertheless, while direct approaches to<br />
assess absolute cancer risks posed by such low-level exposures may elude<br />
us we can nevertheless begin to determine relative cancer risks and<br />
prioritise genotoxic chemicals on the basis <strong>of</strong> experimental determinations<br />
<strong>of</strong> mutagenic potency and estimates <strong>of</strong> target doses resulting from<br />
environmental or occupational exposures. Thus, according to the foregoing<br />
the relative cancer risk posed by a low exposure to a genotoxic agent<br />
would approximate to the number <strong>of</strong> mutations induced per unit target<br />
dose×estimated human target dose. Once relative cancer risks have been
192 MOLECULAR APPROACHES TO ASSESS CANCER RISKS<br />
established, the determination <strong>of</strong> the ‘absolute’ risks for one genotoxic<br />
chemical would permit calculation <strong>of</strong> the ‘absolute’ risks for the others.<br />
Such ‘absolute’ risks will nevertheless vary from population to population<br />
dependent upon variations in promoter pressure.<br />
Determination <strong>of</strong> ‘absolute’ cancer risks<br />
Increments in human mutation caused by low-level exposures to genotoxic<br />
chemicals are essential for risk estimation but cannot be determined<br />
directly (vide supra). Such increments may be estimated using experimental<br />
models. However, unless the variations in background are <strong>of</strong> a very low<br />
order, it is unlikely that even the most sensitive <strong>of</strong> the emerging mutation<br />
assays will permit the measurement <strong>of</strong> small increments <strong>of</strong> mutation at<br />
low, e.g. environmental, exposures. Extrapolation to low doses will be<br />
required and must necessarily be conservative, i.e. linear extrapolation to<br />
the origin.<br />
In addition to ‘high’ dose-low dose extrapolation, it will be necessary to<br />
apply corrections for differences between the model and humans in the<br />
operation <strong>of</strong> systemic factors that govern the relationships between<br />
exposure and mutagenic effect. Estimates <strong>of</strong> target dose in the human<br />
population at risk and in the experimental model compensate for<br />
differences in metabolic and biokinetic factors that determine the<br />
relationships between exposure and the critical dose. In effect, the<br />
determination <strong>of</strong> target dose provides a measure <strong>of</strong> the rates <strong>of</strong> formation <strong>of</strong><br />
the key (primary and critical) chemical lesions leading to mutation. The<br />
final stage in translating the experimentally-determined risk data to<br />
humans is to apply corrections for systemic factors that determine the<br />
progression <strong>of</strong> the key lesions into mutations.<br />
The equivalent radiation dose concept<br />
The principal systemic factors determining the progression <strong>of</strong> key chemical<br />
lesions in DNA into mutations are the rates and fidelities <strong>of</strong> DNA repair<br />
and replication (Wright et al., 1988). Ehrenberg and co-workers have<br />
suggested that the repair <strong>of</strong> primary DNA damage induced by low doses <strong>of</strong><br />
radiation may be proportionate to that induced by low doses <strong>of</strong> genotoxic<br />
chemicals. They have further suggested that the determination <strong>of</strong> the<br />
relative mutagenic effectiveness or potencies <strong>of</strong> radiation and any<br />
particular genotoxic chemical may be <strong>of</strong> value in correcting for species<br />
differences in factors determining the progression <strong>of</strong> primary DNA damage<br />
into mutations. The model proposed by Ehrenberg (1980) is based on the<br />
determination <strong>of</strong> the dose-response curves for the induction <strong>of</strong> the same<br />
mutation in the same experimental system by low target doses <strong>of</strong> the test<br />
chemical and acute -radiation. A consistent ratio between the two curves,
A.S.WRIGHT ET AL. 193<br />
which need not be linear, would indicate proportionality over the low dose<br />
range <strong>of</strong> interest and would permit the mutagenic potency <strong>of</strong> the chemical<br />
to be expressed in terms <strong>of</strong> radiation equivalents, i.e. the number <strong>of</strong> rads<br />
giving the same response or risk as a unit <strong>of</strong> chemical dose (expressed in<br />
terms <strong>of</strong> target dose, e.g. millimolar hour, mM h). The significance <strong>of</strong> such<br />
radiation dose-equivalents hinges on their possible extrapolative value.<br />
Thus, in order to be useful in assisting the translation <strong>of</strong> experimentallydetermined<br />
mutagenicity data, the rad-equivalence value for the test<br />
chemical must have a similar numerical value in both the test system used<br />
to determine mutagenicity and in humans. However, it is improbable that<br />
rad-equivalence values can be directly determined in humans.<br />
Rad-equivalence values for the induction <strong>of</strong> mutations have been<br />
determined for a number <strong>of</strong> intrinsically reactive mon<strong>of</strong>unctional alkylating<br />
agents using a wide range <strong>of</strong> genetic endpoints in a variety <strong>of</strong> biological<br />
systems including bacteria, plants and mammalian species—the latter,<br />
mainly in vitro (Ehrenberg et al., 1974; Ehrenberg, 1976, 1979; Calleman,<br />
1984; Kolman et al., 1989). The rad-equivalence value for a given<br />
alkylating agent was approximately the same (within a factor <strong>of</strong> two) in<br />
each <strong>of</strong> the test systems. On the basis <strong>of</strong> such evidence Calleman et al.<br />
(1978) concluded that there was no reason to presume that a value for radequivalence<br />
established in these disparate systems would differ in humans.<br />
The best studied example is ethylene oxide. Currently a conjoint programme<br />
is underway at the Universities <strong>of</strong> Stockholm and Leiden to determine radequivalence<br />
values in rodents in vivo using a variety <strong>of</strong> endpoints including<br />
the clonal HGPRT mutation assay and induction <strong>of</strong> pre-neoplastic nodules<br />
in rat liver. Preliminary findings are encouraging (Ehrenberg, personal<br />
communication).<br />
Demonstration <strong>of</strong> their extrapolative value would justify application <strong>of</strong><br />
rad-equivalence values to compute small increments in mutation induced in<br />
humans by low exposures (determined as target dose) to genotoxic<br />
chemicals. The determination <strong>of</strong> these increments is the basis <strong>of</strong> the risk<br />
model (vide supra) in which the increment in cancer risk due to a particular<br />
chemical in a population is viewed as approximating to the increment in<br />
mutation induced by the chemical. However, the fact that risk coefficients<br />
have not yet been established for radiation-induced mutations in humans<br />
precludes applications <strong>of</strong> rad-equivalence values to estimate mutational<br />
risks, e.g. small increments in mutation, in humans. Of course, Ehrenberg<br />
realised that a genotoxic chemical(s) may prove to be superior to radiation<br />
as a reference standard for estimating mutational risks. Indeed, the use <strong>of</strong><br />
chemicals that are representative <strong>of</strong> classes <strong>of</strong> genotoxic chemicals, repair<br />
pathways, etc. is envisaged in this developing ‘equivalence’ strategy<br />
(Törnqvist and Osterman-Golkar, 1991). However, at the time the original<br />
strategy was formulated there was and still is no reliable data relating low<br />
level exposure to a genotoxic chemical and the attendant risk <strong>of</strong> mutation
194 MOLECULAR APPROACHES TO ASSESS CANCER RISKS<br />
or cancer in human populations. In contrast, risk coefficients have been<br />
established for the induction <strong>of</strong> cancer by low levels <strong>of</strong> -radiation in<br />
human populations within, approximately, a factor <strong>of</strong> 2.<br />
According to the basic presumptions (vide supra), the increased cancer<br />
risks in a human population caused by low-level exposures to radiation or<br />
to a genotoxic chemical are predominantly due to the mutagenic<br />
propensities <strong>of</strong> these genotoxic factors. It would, therefore, seem<br />
reasonable to suggest that cells that had been initiated by exposure to<br />
radiation or to genotoxic chemicals or combinations <strong>of</strong> factors (vide supra)<br />
would, nevertheless, be subject to the same general promoter and<br />
modulating influences acting on the population. On this basis,<br />
experimentally-determined rad-equivalence values for genotoxic chemicals<br />
may be used to convert target doses <strong>of</strong> the chemicals (determined in human<br />
populations receiving low-level exposures) into the equivalent doses <strong>of</strong><br />
radiation (for the induction <strong>of</strong> mutation). The respective cancer risks<br />
associated with any particular monitored target dose <strong>of</strong> a genotoxic<br />
chemical may then be obtained by direct reference to the human cancer risk<br />
coefficients established for radiation.<br />
References<br />
ALBRECHT, W. and NEUMANN, H.-G., 1985, Biomonitoring <strong>of</strong> aniline and<br />
nitrobenzene. Haemoglobin binding in rats and analysis <strong>of</strong> adducts, Arch.<br />
Toxicol, 57, 1–5.<br />
AMES, B.N., 1989, Environmental pollution and the causes <strong>of</strong> human cancer: six<br />
errors, in DeVita, V.T., Jr, Hellman, S. and Rosenberg, S.A. (Eds) Important<br />
Advances in Oncology, pp. 237–47, Philadelphia, PA: Lippincott.<br />
AMES, B.N., DURSTON, W.E., YAMASAKI, E. and LEE, F.E., 1973, Carcinogens<br />
are mutagens: a simple test system combining liver homogenates for activation<br />
and bacteria for detection, Proc. Nat. Acad. Sci. USA, 70, 2281–5.<br />
CALLEMAN, C.J., 1984, Haemoglobin as a dose monitor and its application to<br />
the risk estimation <strong>of</strong> ethylene oxide, PhD Thesis, p. 25, Stockholm: University<br />
<strong>of</strong> Stockholm.<br />
CALLEMAN, C.J., EHRENBERG, L., JANSSON, B., OSTERMAN-GOLKAR, S.,<br />
SEGERBÄCK, D., SVENSSON, K. and WACHTMEISTER, C.A., 1978,<br />
Monitor ing and risk assessment by means <strong>of</strong> alkyl groups in haemoglobin in<br />
persons occupationally exposed to ethylene oxide, J. Environm. Pathol. Tox.,<br />
2, 427–42.<br />
COOPER, D.P., GRIFFIN, K.A. and POVEY, A.C., 1992, Immunoaffinity<br />
purification combined with 32 P-postlabelling for the detection <strong>of</strong> O 6 -<br />
methylguanine in DNA from human tissues, Carcinogenesis, 13, 469–75.<br />
EHRENBERG, L., 1974, Genotoxicity <strong>of</strong> environmental chemicals, Acta. Biol.<br />
Yugosl., Ser. F Genetika, 6, 367–98.<br />
EHRENBERG, L., 1976, Methods <strong>of</strong> Comparing Effects <strong>of</strong> Radiation and<br />
Chemicals, Brighton IAEA Consultant Meeting.
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EHRENBERG, L., 1979, Risk assessment <strong>of</strong> ethylene oxide and other compounds,<br />
in McElheny, V.K. and Abrahamson, S. (Eds) Assessing Chemical Mutagens:<br />
The Risk to Humans (Banbury Report 1), pp. 157–90, Cold Spring Harbour,<br />
New York, CSH Press.<br />
EHRENBERG, L., 1980, Purposes and methods <strong>of</strong> comparing risks <strong>of</strong> radiation<br />
and chemicals, in Radiobiological Equivalents <strong>of</strong> Chemical Pollutants, pp. 11–<br />
36, Vienna: International Atomic Energy Agency.<br />
EHRENBERG, L., HIESCHE, K.D., OSTERMAN-GOLKAR, S. and WENNBERG,<br />
I., 1974, Evaluation <strong>of</strong> genetic risks <strong>of</strong> alkylating agents: tissue doses in the<br />
mouse from air contaminated with ethylene oxide, Mutat. Res., 24, 83–103.<br />
EHRENBERG, L. and OSTERMAN-GOLKAR, S., 1980, Alkylation <strong>of</strong><br />
macromolecules for detecting mutagenic agents, Teratogen., Carcinogen.<br />
Mutagen., 1, 105–27.<br />
FARMER, P.B., 1991, Analytical approaches for the determination <strong>of</strong><br />
proteincarcinogen adducts using mass spectrometry, in Groopman, J.D. and<br />
Skipper, P. L. (Eds) Molecular Dosimetry and Human Cancer: Analytical,<br />
Epidemiological and Social Considerations, pp. 189–210, Boca Raton: CRC<br />
Press.<br />
GREEN, L.C., SKIPPER, P.L., TURESKY, R.J., BRYANT, M.S. and<br />
TANNENBAUM, S.R., 1984, In vivo dosimetry <strong>of</strong> 4-aminobiphenyl in rats via<br />
a cysteine adduct in haemoglobin, Cancer Res., 44, 4254–9.<br />
KOLMAN, A., NASLUND, M., OSTERMAN-GOLKAR, S., SCALIA-TOMBA,<br />
G.P. and MEYER, A.L., 1989, Comparative studies <strong>of</strong> in vitro transformation<br />
by ethylene oxide and gamma-radiation <strong>of</strong> cells, Mutagenesis, 4, 58–61.<br />
Lu, A-L. and Hsu, I-C., 1992, Detection <strong>of</strong> single DNA base mutations with<br />
mismatch repair enzymes, Genomics, 14, 249–55.<br />
OSTERMAN-GOLKAR, S., EHRENBERG, L., SEGERBÄCK, D. and<br />
HALLSTROM, I., 1976, Evaluation <strong>of</strong> genetic risks <strong>of</strong> alkylating agents. II.<br />
Haemoglobin as a dose monitor, Mutat. Res., 34, 1–10.<br />
RANDERATH, K., REDDY, M.V. and GUPTA, R.C., 1981, 32 P-labeling test for<br />
DNA damage, Proc. Natl Acad. Sci. USA, 78, 6126–9.<br />
SKIPPER, P. L and NAYLOR, S., 1991, Mass spectrometric analysis <strong>of</strong><br />
proteincarcinogen adducts, in Garner, R.C, Farmer, P.B., Steel, G.T. and<br />
Wright, A.S. (Eds) Human Carcinogen Exposure: Biomonitoring and Risk<br />
Assessment, pp. 61–8, Oxford: Oxford University Press.<br />
THILLY, W.G., 1991, Mutational spectrometry: opportunities and limitations in<br />
human risk assessment, in Garner, R.C, Farmer, P.B., Steel, G.T. and Wright,<br />
A. S. (Eds) Human Carcinogen Exposure: Biomonitoring and Risk<br />
Assessment, pp. 127–33, Oxford: Oxford University Press.<br />
TÖRNQVIST, M. and OSTERMAN-GOLKAR, S., 1991, Monitoring <strong>of</strong> in vivo<br />
dose by macromolecular adducts: usefulness in risk estimation, in Groopman,<br />
J.D. and Skipper, P.L. (Eds) Molecular Dosimetry and Human Cancer:<br />
Analytical, Epidemi ological and Social Considerations, pp. 89–102, Boca<br />
Raton: CRC Press.<br />
TÖRNQVIST, M., MOWRER, J., JENSEN, S. and EHRENBERG, L., 1986a,<br />
Monitoring <strong>of</strong> environmental cancer initiators through haemoglobin adducts<br />
by a modified Edman degradation method, Analyt. Biochem., 154, 255–66.
196 MOLECULAR APPROACHES TO ASSESS CANCER RISKS<br />
TÖRNQVIST, M., OSTERMAN-GOLKAR, S., KAUTIAINEN, A., JENSEN, S.,<br />
FARMER P.B. and EHRENBERG, L., 1986b, Tissue doses <strong>of</strong> ethylene oxide<br />
in cigarette smokers determined from adduct levels in haemoglobin,<br />
Carcinogenesis, 7, 1519–21.<br />
WRAITH, M.J., WATSON, W.P., EADSFORTH, C.V., VAN SITTERT, N.J.,<br />
TÖRNQVIST, M. and WRIGHT, A.S., 1988, An immunoassay for monitoring<br />
human exposure to ethylene oxide in Bartsch, H., Hemminki, K., and O’Neill,<br />
I.K. (Eds) Methods for Detecting DNA Damaging Agents in Humans:<br />
Applications in Cancer Epidemiology and Prevention, IARC Scientific<br />
Publications No. 89, pp. 271–4, Lyon, France: International Agency for<br />
Research on Cancer.<br />
WRIGHT, A.S., 1991, Emerging strategies for the determination <strong>of</strong> human<br />
carcinogens, detection, identification, exposure monitoring and risk<br />
evaluation, in Garner, R.C, Farmer, P.B., Steel, G.T. and Wright, A.S. (Eds)<br />
Human Carcinogen Exposure: Biomonitoring and Risk Assessment, pp. 3–23,<br />
Oxford: Oxford University Press.<br />
WRIGHT, A.S., BRADSHAW, T.K. and WATSON, W.P., 1988, Prospective<br />
detection and assessment <strong>of</strong> genotoxic hazards: a critical appreciation <strong>of</strong> the<br />
contribution <strong>of</strong> L.G.Ehrenberg, in Bartsch, H., Hemminki, K. and O’Neill, I.K.<br />
(Eds) Methods for Detecting DNA Damaging Agents in Humans: Applications<br />
in Cancer Epidemiology and Prevention, IARC Scientific Publications No. 89,<br />
pp. 237–47, Lyon, France: International Agency for Research on Cancer.
15<br />
Evaluation <strong>of</strong> Toxicity to the Immune System<br />
HANS-WERNER VOHR<br />
Bayer AG, Wuppertal<br />
Introduction<br />
A number <strong>of</strong> years ago the new field <strong>of</strong> immunotoxicology was established.<br />
Very early on, calls were heard from various sides demanding that the<br />
development <strong>of</strong> new chemicals should also take into account the influence<br />
<strong>of</strong> these substances on the immune system (Dean, 1979; Luster et al., 1988;<br />
Trizio, et al., 1988). These demands led on the one hand to the initiation <strong>of</strong><br />
a number <strong>of</strong> investigations and collaborative studies (ICICIS; BGA; US-<br />
NTP), on the other hand to thoughts by the authorities and industry on the<br />
introduction <strong>of</strong> guidelines (US-EPA, 1982, 1990; Sjoblad, 1988; ECETOC,<br />
1990; UK-DOH, 1991; Hinton, 1992; OECD, 1992ab.<br />
If we define immunotoxicology as the science <strong>of</strong> adverse effects <strong>of</strong><br />
substances on the immune system we can say further that these side-effects<br />
can lead to either immunopotentiation or immunosuppression. The former<br />
can lead to induction <strong>of</strong> autoimmune reactions and to Type I-IV<br />
hypersensitivity reactions, the latter to reduced resistance to infection,<br />
development <strong>of</strong> cancer and also to autoimmune phenomena.<br />
On the basis <strong>of</strong> this definition, immunotoxicological investigations have<br />
already been carried out for years during the development <strong>of</strong> substances;<br />
namely with respect to DTH reactions (Type IV) in the guinea pig (Bühler,<br />
1965; Magnusson and Kligman, 1969). As an alternative to these tests in<br />
the guinea pig, the so-called local lymph node assay (LLNA) in the mouse<br />
according to Kimber et al. (1989) was developed and validated and has<br />
meanwhile been adopted as alternative test in the OECD guidelines (OECD,<br />
1992a, b; Botham et al., 1991).<br />
The development or selection <strong>of</strong> suitable tests for immunotoxicological<br />
screening and thus for incorporation in guidelines presents considerable<br />
problems. Most <strong>of</strong> the tests which have been proposed for<br />
immunotoxicological investigations and most knowledge and experience in<br />
immunology are based on mouse models. The standard animal in the early<br />
phase <strong>of</strong> toxicological testing, however, is the rat. Transference <strong>of</strong> the tests<br />
is not always easy, partly because <strong>of</strong> lack <strong>of</strong> suitable reagents.
198 EVALUATION OF TOXICITY TO THE IMMUNE SYSTEM<br />
The next problem is to find which tests can suitably be used—simply and<br />
without having to treat additional animals—for reliable identification <strong>of</strong><br />
interactions with the immune system. Another question which has not yet<br />
been solved is relating to the dosages and the changes in immunological<br />
parameters which are still tolerable and at which times these should be<br />
determined.<br />
National and international collaborative studies<br />
Most <strong>of</strong> the guideline drafts favour a two-tiered to three-tiered approach<br />
for the screening <strong>of</strong> immunotoxic side effects, with the greatest disparities<br />
in respect <strong>of</strong> the proposals for tier I tests, ranging from just organ weights<br />
and a little more emphasis on the histology <strong>of</strong> the lymphatic organs, to a<br />
series <strong>of</strong> elaborate, supplementary function tests including, in some cases,<br />
satellite groups.<br />
Any discussion about basic tests is hampered by a lack <strong>of</strong> data from<br />
routine toxicological and/or epidemiological studies although a few years<br />
ago a number <strong>of</strong> collaborative studies were initiated, namely: ICICIS<br />
(international), US-NTP (international, USA (two)), BGA (Federal German<br />
Health Office, Germany), GEVI (France). The aim <strong>of</strong> all these studies is to<br />
check on different histological parameters and functional tests for their<br />
possibility to flag an immunotoxic potential <strong>of</strong> a compound in a routine 14<br />
or 28 day study (rats) in an interlaboratory trial. Two immunosuppressive<br />
standards (azathioprine and/or cyclosporin A) have been used so far.<br />
Although the experimental phase is finished the evaluation <strong>of</strong> the data is<br />
still underway.<br />
Table 15.1 summarises the immunotoxicological experiments and<br />
collaborative studies currently in progress. With a few exceptions all<br />
investigations are based on a 28-day gavage study in rats. These basic tests<br />
were supplemented by extended histopathology and functional tests.<br />
ICICIS<br />
The first substance to be investigated in the international collaborative<br />
study was azathioprine. However this first attempt suffered from lack <strong>of</strong><br />
harmonisation between the models used by the 28 participants world-wide.<br />
This made the comparability and the evaluation <strong>of</strong> the results particularly<br />
problematic.<br />
The second substance which was tested by ICICIS in a more restrictive<br />
design with slightly fewer participants was cyclosporin A. The<br />
experimental section as well as the first evaluation <strong>of</strong> the results has now<br />
been completed.<br />
As the final evaluation—including statistics—<strong>of</strong> ICICIS is likely to take<br />
some time (years?) it will not be discussed further at this point.
Table 15.1 Immunotoxicology collaborative studies<br />
Fischer 344 study (Kimber-White)<br />
H.-W.VOHR 199<br />
An other interesting approach was done by Kimber-White and colleagues.<br />
In this study the duration <strong>of</strong> treatment was 14 days, Fischer 344 rats were<br />
used as experimental animals. Apart from the usual parameters the plaque<br />
assay (PFCA), mitogenic stimulation (ConA, LPS) and NK activity were<br />
measured.
200 EVALUATION OF TOXICITY TO THE IMMUNE SYSTEM<br />
No effects were found with respect to organ weights or cell counts<br />
analysis but there were marked effects in the PFCA. Here there was a good<br />
correlation between the laboratories and clearly dose-dependent effects<br />
were seen. Mitogenic stimulation was not and NK activity only slightly<br />
affected.<br />
Results <strong>of</strong> the BGA collaborative study<br />
In order to put the discussion on a somewhat sounder footing it is<br />
imperative to test the various models for detection <strong>of</strong> immunotoxicological<br />
potential in practice. For this reason Bayer AG is taking part in a<br />
collaborative trial initiated by the German Federal Health Office in Berlin<br />
(BGA study). In this collaborative study standard immunotoxic substances<br />
are investigated in parallel under restrictive conditions in several<br />
laboratories. On the other hand Bayer AG has introduced a set <strong>of</strong><br />
functional immunological tests into the routine toxicological testing <strong>of</strong><br />
agrochemicals in rats in order to test the informative value <strong>of</strong> these<br />
parameters in practice (Vohr, 1995).<br />
In the course <strong>of</strong> this collaborative study cyclosporin A was investigated<br />
as the first substance in a very well harmonised design. One reason for<br />
choosing cyclosporin A was to permit comparison with the ICICIS results.<br />
The study was based on OECD guideline 407 which was supplemented by<br />
a number <strong>of</strong> histopathological, haematological, clinical-chemical and<br />
functional parameters. The final evaluation, including pathology and<br />
statistics will take a few more months. Nevertheless a first look at the data<br />
showed that there were no marked effects with respect to either organ<br />
weights (lymphatic organs) or blood cells. On the other hand dose-related<br />
effects were found for many parameters in the functional tests.<br />
Examples are shown <strong>of</strong> these parameters and their changes at the various<br />
dosages. Both sexes show dose-related changes from the lowest dosage (1<br />
mg kg −1 ) upwards with respect to the surface markers <strong>of</strong> the<br />
immunocompetent cells. The PFCA and the measurements <strong>of</strong> IgA<br />
antibodies in serum also proved to be sensitive parameters.<br />
In immunotoxic investigations based on data obtained in rats treated<br />
with test substances (28-day test) we found that secondary influences <strong>of</strong>ten<br />
occur and that occasionally one single parameter is changed at the highest<br />
dosage. Apart from these non-specific effects, dose-related reactions were<br />
observed from the lowest or middle dose upwards (for example in the case<br />
<strong>of</strong> cytostatic substances). Such genuinely immunotoxic compounds were<br />
reliably identified by surface markers (like CD4/CD45R or PanB) and<br />
changes in the serum Ig-titres. Changes in other parameters such as cell<br />
counts and macrophage activity verified these findings.
Findings <strong>of</strong> the US-NTP study (Luster)<br />
The US-NTP study investigated 51 substances, 35 <strong>of</strong> which were declared<br />
immunotoxic, in a comprehensive test battery in mice for changes in<br />
functional parameters after 28-day administration <strong>of</strong> the substances. The<br />
correlations <strong>of</strong> each <strong>of</strong> these parameters with the given classification and<br />
with the results <strong>of</strong> host-resistance studies were calculated. The correlations<br />
after combinations <strong>of</strong> individual tests were also calculated (Luster et al.,<br />
1992, 1993).<br />
The conclusion drawn from these investigations was that the<br />
immunotoxic potential <strong>of</strong> a substance can be determined relatively reliably<br />
by combination <strong>of</strong> 2–3 specific tests. The most powerful tests proposed by<br />
Luster et al. for such a combination include surface markers, NK test and<br />
PFCA. Serum titres <strong>of</strong> Ig —particularly IgA—were unfortunately not<br />
determined.<br />
With regard to the correlation with host resistance (HR) results it was<br />
found that if effects were shown in the HR model there were always effects<br />
on functional parameters, too. There were, however, also cases in which<br />
there were effects in the functional tests although the HR studies were<br />
negative. Although these investigations were carried out on mice and the<br />
choice and classification <strong>of</strong> the substances are not entirely undisputed,<br />
these findings are nevertheless confirmed by our own experience.<br />
Discussion and prospect<br />
H.-W.VOHR 201<br />
It is undoubtedly too early to make any judgement. However, it appears<br />
that apart from the histology—particularly the immuno-histology—a few<br />
additional parameters such as analysis <strong>of</strong> surface markers <strong>of</strong> subpopulations<br />
and serum titre assays <strong>of</strong> IgG, IgM and IgA are sufficient as screening<br />
indicators to show the possible immunotoxic potential <strong>of</strong> a substance. One<br />
<strong>of</strong> these is the PFCA, which presupposes, however, that satellite groups are<br />
used or that the authorities accept injection <strong>of</strong> SRBC as GLP treatment.<br />
Positive findings in a combination <strong>of</strong> these tests should then occasion<br />
further elucidation <strong>of</strong> the immunotoxic potential.<br />
In summary it can be concluded from the currently available results <strong>of</strong> the<br />
collaborative studies that the following criteria must be fulfilled if a<br />
substance is to be characterised as possibly immunotoxic. The substance<br />
must:<br />
1. Induce significant dose-related changes in one <strong>of</strong> the effective tests<br />
listed above, or<br />
2. induce significant changes in the highest dose group in a combination<br />
<strong>of</strong> 2–3 <strong>of</strong> these tests.
202 EVALUATION OF TOXICITY TO THE IMMUNE SYSTEM<br />
But also findings in lymphoid organs which are remarkable with respect to<br />
quality, severity or quantity <strong>of</strong> changes should be <strong>of</strong> sufficient relevance to<br />
warrant further assessment.<br />
Finally I would like to point out the problems involved in the evaluation<br />
<strong>of</strong> immunological changes.<br />
• In contrast to other data obtained in routine toxicology, historical data<br />
and routine experience with respect to the functional tests are so far<br />
almost entirely lacking. On account <strong>of</strong> this paucity <strong>of</strong> data a decision as<br />
to the right effect level for the functional tests can only be made on the<br />
basis <strong>of</strong> subjective judgement at present. The discussion <strong>of</strong> uniform<br />
criteria is only just beginning.<br />
• If effects only occur secondarily, e.g. on account <strong>of</strong> inflammatory<br />
processes, they are not specifically immunotoxic. It must be discussed<br />
whether such effects on the immune system, which is only doing its<br />
normal job in these cases, can be classed as immunotoxic at all.<br />
• The immune system can show an ‘oscillating’ response to a substance. A<br />
substance may have a stimulating effect in a low dose range and an<br />
inhibitory effect in a high range or vice versa. Such reactions are also<br />
dose-related effects.<br />
• Most immunotoxic investigations have so far used known<br />
immunosuppressive drugs. There are, however, little data—particularly<br />
from rat studies— on immunostimulant substances. Since the other<br />
unwanted side-effect apart from immunosuppression is immune<br />
potentiation, future collaborative investigations must urgently<br />
concentrate on such substances. Before this, no final evaluation and thus<br />
no recommendations on relevant tests can be made.<br />
For pharmaceuticals (Hinton, 1992), pesticides (US-EPA, 1990, 1993) and<br />
veterinary medicinal chemicals (EEC, 1991) final drafts or notes for<br />
guidance for the screening <strong>of</strong> the immunotoxic potential <strong>of</strong> a compound<br />
already exist. These draft proposals for immunotoxicity parameters for<br />
incorporation into new guidelines are shown in Tables 15.2 (FDA) and<br />
15.3 (US-EPA).<br />
For industrial chemicals advanced screening <strong>of</strong> lymphoid organs also<br />
with respect to functional parameters had been expected to be incorporated<br />
into the adopted OECD guideline No. 407 (1992). But recommendations<br />
made by van Loveren and Vos (1992) have not yet been taken into<br />
consideration for this update <strong>of</strong> OECD guideline 407. The proposal <strong>of</strong><br />
these authors recommended more histopathology (gut associated lymphoid<br />
tissue), measurement <strong>of</strong> serum immunoglobulins, bone marrow cellularity,<br />
cyt<strong>of</strong>luorimetry <strong>of</strong> spleen cells and measurement <strong>of</strong> NK cell activity.<br />
A task force ‘Immunotoxicology’ initiated by ECETOC has put forward<br />
proposals on hazard identification and risk assessment <strong>of</strong> immunotoxic
Table 15.2 Summary <strong>of</strong> immunotoxicity testing recommendations for direct food additives<br />
H.-W.VOHR 203<br />
Abbreviations: CBC=complete blood cell count; WBC=white blood cell count; Ig=Immunoglobin; NK=natural killer;<br />
IL-2=interleukin 2; SRBC=sheep red blood cells; TNP-LPS=trinitrophenol lipopolysaccharide.<br />
* Recommended for inclusion in basic testing.<br />
potential <strong>of</strong> a compound on the basis <strong>of</strong> the routine 28 day treatment <strong>of</strong><br />
rats (Basketter et al., 1994; ECETOC, 1994). A central point in these
204 EVALUATION OF TOXICITY TO THE IMMUNE SYSTEM<br />
Table 15.3 Proposed amendment to the subdivision F guideline requirements to<br />
provide for an evaluation <strong>of</strong> the immunotoxicity <strong>of</strong> chemical pesticides<br />
proposals is a flow diagram for the evaluation <strong>of</strong> results from this basic<br />
study for hazard identification and conclusions drawn from it. This flow<br />
diagram, shown in Figure 15.1, could be helpful for the evaluation <strong>of</strong><br />
results obtained from incorporation <strong>of</strong> a basic immunotoxicity test battery<br />
into studies with routinely treated animals.<br />
References<br />
BASKETTER, D. et al., 1994, The identification with sensitising or<br />
immunosuppressive properties in routine toxicology, Food Chem. Toxicol. (in<br />
press).<br />
BOTHAM, P.A. et al., 1991, Skin sensitization—a critical review <strong>of</strong> predictive test<br />
methods in animals and men Food Chem. Toxicol, 29, 275–86.<br />
BÜHLER, E.V., 1965, Delayed contact hypersensitivity in the guinea pig, Arch.<br />
Dermatol. 91, 171.<br />
DEAN, J.H., PADARATHSINGH, M.L. and JERRELS, T.R., 1979, Assessment <strong>of</strong><br />
immunobiological effects induced by chemicals, drugs or food additives. I. Tier<br />
testing and screening approach, Drug Chem. Toxicol. 2, 5–17.<br />
ECETOC, 1990, Skin Sensitization Testing, Monograph No. 14, Brussels.
Figure 15.1 Flow diagram (taken from the ECETOC monograph:<br />
Immunotoxicity).<br />
H.-W.VOHR 205<br />
ECETOC, 1994, Immunotoxicity: hazard identification and risk assessment,<br />
Monograph No. 21, Brussels.<br />
EEC, 1991, Annex to directive 81/852/EEC—Note <strong>of</strong> guidance, 22.
206 EVALUATION OF TOXICITY TO THE IMMUNE SYSTEM<br />
HINTON, D.M., 1992, Testing guidelines for evaluation <strong>of</strong> the immunotoxic<br />
potential <strong>of</strong> direct food additives, Crit. Rev. Food Sci. Nutrit, 32, 173–90.<br />
KIMBER, I., HILTON, J. and WEISENBERGER, C. 1989, The murine local lymph<br />
node assay for identification <strong>of</strong> contact allergens: a preliminary evaluation <strong>of</strong><br />
in situ measurement <strong>of</strong> lymphocyte proliferation, Contact Dermatit. 21, 215–<br />
20.<br />
LUSTER, M.I. et al., 1988, Development <strong>of</strong> a battery to assess chemical-induced<br />
immunotoxicity: National <strong>Toxicology</strong> Program’s guidelines for<br />
immunotoxicity evaluation in mice, Fundam. Appl. Toxicol. 10, 2–19.<br />
LUSTER, M.I. et al., 1992, Risk assessment in immuno-toxicology. I. Sensitivity<br />
and predictability <strong>of</strong> immune tests, Fundam. Appl. Toxicol., 18, 200–10.<br />
LUSTER, M.I. et al., 1993, Risk assessment in immuno-toxicology. II. Relationship<br />
between immune and host resistance tests, Fundam. Appl. Toxicol., 21, 71–82.<br />
MAGNUSSON, B. and KLIGMANN, A.M., 1969, The identification <strong>of</strong> contact<br />
allergens by animal assay. The guinea pig maximisation test, J. Invest.<br />
Dermatol., 52, 268.<br />
OECD, 1992a, Organisation for Economic Cooperation and Development,<br />
Guidelines for testing <strong>of</strong> chemicals No. 407, adopted 12 July, 1992.<br />
OECD, 1992b, Organisation for Economic Cooperation and Development,<br />
Guidelines for testing <strong>of</strong> chemicals—skin sensitisation, No. 406, adopted 17<br />
July, 1992.<br />
SJOBLAD, R., 1988, Potential future requirements for immunotoxicology testing<br />
<strong>of</strong> pesticides, Toxicol. Indust. Hlth, 4, 391.<br />
TRIZIO, D. et al., 1988, Identification <strong>of</strong> immunotoxic effects <strong>of</strong> chemicals and<br />
assessment <strong>of</strong> their relevance to man. Food Chem. Toxic., 26, 527–39.<br />
UK Department <strong>of</strong> Health, 1991, Proposed to update OECD Guideline 407.<br />
US-EPA, 1982, Code <strong>of</strong> Federal Regulations, Washington DC, 152–18, 152–24 and<br />
158–165.<br />
US-EPA, 1990, Draft immunotoxicity study screen for testing chemical pesticides.<br />
VAN LOVEREN, H. and Vos, J.G., 1992, Evaluation <strong>of</strong> OECD Guideline 407 for<br />
assessment <strong>of</strong> toxicity <strong>of</strong> chemicals with respect to potential adverse effects to<br />
the immune system. RIVM Report No. 158801001, Bilthoven: National<br />
Institute <strong>of</strong> Public Health and Environmental Protection.<br />
VOHR, H.-W., 1995, Experiences with an advanced screening procedure for the<br />
identification <strong>of</strong> chemicals with an immunotoxic potential in routine<br />
toxicology (a position paper). <strong>Toxicology</strong>, (in press),
16<br />
New Strategies: the Use <strong>of</strong> Long-term Cultures <strong>of</strong><br />
Hepatocytes in Toxicity Testing and Metabolism<br />
Studies <strong>of</strong> Chemical Products Other than<br />
Pharmaceuticals<br />
VERA ROGIERS, 1 MAY AKRAWI, 2 SANDRA COECKE, 1<br />
YVES VANDENBERGHE, 1 ELIZABETH SHEPHARD, 2 IAN<br />
PHILLIPS 3 and ANTOINE VERCRUYSSE 1<br />
1 Vrije Universiteit Brussel, Brussels; 2 University College<br />
London, London; 3 University <strong>of</strong> London, London<br />
Introduction: metabolism and toxicity <strong>of</strong> chemical<br />
products are closely linked<br />
Lipophilic compounds are metabolized in the liver by phase 1 and phase 2<br />
reactions into more polar, more hydrophilic metabolites, which are usually<br />
less biologically active. Bioactivation, however, may occur, forming toxic<br />
species by phase 1, cytochrome P450 (CYP) dependent oxidation (e.g.<br />
epoxidation <strong>of</strong> C=C to reactive epoxide intermediates (Guengerich et al.<br />
1991), CYP dependent reduction (e.g. dehalogenation <strong>of</strong> CCl 4 toa free<br />
radical intermediate) (Timbrell, 1993) or even by phase 2 reactions (e.g.<br />
reactive episulphonium ion formation by glutathione conjugation <strong>of</strong><br />
dibromoethane) (Van Bladeren, 1988; Coles and Ketterer, 1990; Timbrell,<br />
1993).<br />
The major process involved in the bioactivation <strong>of</strong> chemical carcinogens<br />
is their oxidation catalyzed by CYP enzymes. Thirty or more different CYP<br />
forms exist within each animal species (Nebert et al., 1989), each with at<br />
least some distinct elements <strong>of</strong> catalytic specificity and regulation. The role<br />
<strong>of</strong> some <strong>of</strong> these CYPs in the activation and detoxication <strong>of</strong> chemical<br />
carcinogens has already been determined. For example:<br />
– CYP2E1 is a major catalyst involved in the oxidation <strong>of</strong> benzene,<br />
styrene, CCl 4, CHCl 3, ethylene dichloride, vinylchloride, acrylonitrile,<br />
vinyl carbamate (Guengerich et al., 1991), ethanol (Perrot et al., 1989),<br />
dialkylnitrosamines (Yoo et al., 1988), isobutene (Cornet et al., 1991)<br />
and some other small molecules.<br />
– CYP1A1 is involved in the oxidation <strong>of</strong> polycyclic aromatic<br />
hydrocarbons (Shimada et al., 1989b).<br />
– CYP1A2 activates arylamines (Shimada et al., 1989a).
208 USE OF LONG-TERM CULTURES OF HEPATOCYTES IN TOXICITY TESTING<br />
– CYP3A4 is a major catalyst in the activation <strong>of</strong> aflatoxins, pyrrolizidine<br />
alkaloids and polycyclic hydrocarbon dihydrodiols (Shimada et al.,<br />
1989a; Shimada and Guengerich, 1989).<br />
The balance between the rates <strong>of</strong> formation <strong>of</strong> reactive metabolites and<br />
detoxication will greatly determine the potential toxic response <strong>of</strong> a<br />
chemical. Variables, known to affect normal biotransformation, such as<br />
enzyme induction and inhibition can also change the bioactivation rate <strong>of</strong><br />
chemicals. A clear example is the metabolism <strong>of</strong> halogenated biphenyls<br />
after treatment with arochlor 1254 (Borlakoglu and Wilkins, 1993).<br />
Consequently, the toxicity <strong>of</strong> chemical products <strong>of</strong>ten depends upon their<br />
specific biotransformation, the presence, absence, induction and inhibition<br />
<strong>of</strong> specific phase 1 and phase 2 enzymes involved in their metabolism.<br />
Towards an in vitro approach <strong>of</strong> risk assessment<br />
Nearly all toxicological studies on chemical products, including industrial<br />
chemicals, agrochemicals, pharmaceuticals, additives, materials in contact<br />
with food and cosmetics, have been carried out in vivo using experimental<br />
animals, in particular small vertebrates (News and Views, 1993).<br />
Important scientific, technological, ethical and economic considerations,<br />
however, justify the actual search for in vitro alternatives, replacing or<br />
improving existing in vivo methods and reducing the number <strong>of</strong> animals<br />
involved (Frazier and Goldberg, 1990; Roberfroid, 1991).<br />
Since hepatocytes can be isolated from different species (Guguen-<br />
Guillouzo et al., 1982; Green et al., 1986; van’t Klooster et al., 1992)<br />
including man (Guguen-Guillouzo et al., 1982; Rogiers, 1993), they can<br />
represent a powerful tool for short-term risk assessment studies when used<br />
as suspensions <strong>of</strong> freshly isolated cells (up to 3–4 h) or as short-term<br />
cultures (up to 2 days) (Klaassen and Stacey, 1982; Guillouzo, 1986). They<br />
can be useful in biotransformation, cytotoxicity, hepatotoxicity and<br />
genotoxicity studies, in species selection and in mechanistic studies<br />
(Blaauboer, 1994). Long-term cultures <strong>of</strong> hepatocytes (up to several weeks)<br />
represent a somewhat different approach in risk assessment. Such systems<br />
are <strong>of</strong> interest for the assessment <strong>of</strong> long-term toxicity <strong>of</strong> xenobiotics<br />
including the occurrence <strong>of</strong> enzyme induction, the effects on xenobiotic<br />
biotransformation, lipid peroxidation, accumulation <strong>of</strong> triglycerides,<br />
changes in glutathione content, interaction between compounds and the<br />
hepatoprotection afforded by certain molecules. Consequently, long-term<br />
cultures have been particularly applied in the development <strong>of</strong><br />
pharmaceuticals, in pharmaco-toxicological studies (Guillouzo, 1986,<br />
1992; Rogiers and Vercruysse, 1993; Skett, 1995). As far as chemical<br />
products other than pharmaceuticals and in particular industrial chemicals<br />
and agrochemicals are concerned, the practical needs during development
are different. The compounds brought onto the European market are<br />
labelled and need to fulfill only the legal demands <strong>of</strong> the specific category<br />
to which they belong.<br />
Testing is only obligatory when they reach the market and the type <strong>of</strong><br />
tests needed per category <strong>of</strong> chemicals is clearly outlined. In Europe legal<br />
categories consist <strong>of</strong> dangerous compounds (88/379/EEC, 93/18/EEC), 1a,b<br />
phytopharmaceuticals (91/414/EEC, 93/71/EEC), 2 biocides (93/C239/03), 3<br />
cosmetics (76/ 768/EEC, 93/35/5/EEC) 4a,b and food additives. Of the latter<br />
category the most comprehensive surveys are carried out by the Joint<br />
Expert Committee on Food Additives (JECFA) <strong>of</strong> the World Health<br />
Organization and the Food and Agriculture Organization <strong>of</strong> the United<br />
Nations (Conning, 1993).<br />
Less sophisticated in vitro studies have been performed on industrial<br />
chemicals and agrochemicals than is the case for pharmaceuticals. The only<br />
field in which isolated hepatocytes have already been incorporated into<br />
routine screening <strong>of</strong> industrial chemicals for regulatory purposes is in<br />
genotoxicity testing (Swierenga et al., 1991). The potentialities, however,<br />
<strong>of</strong> in vitro testing for these compounds, in particular <strong>of</strong> the use <strong>of</strong> long-term<br />
cultures, has not yet been explored in depth, although induction, inhibition,<br />
biotransformation, chronic toxicity, interaction between chemicals and<br />
mechanistic studies are <strong>of</strong> great interest for these compounds too.<br />
Human exposure to chemical products such as pesticides, eventually<br />
reaching the food chain as residues or <strong>of</strong> potential risk for workers and<br />
operators spraying the fields, is such an interesting research area. For<br />
example, Alachlor ® , a herbicide, <strong>of</strong> which millions <strong>of</strong> tons are used per<br />
year, has been classified as a potential human carcinogen (Leslie et al.,<br />
1989) because <strong>of</strong> tumour formation in rats and DNA damage observed in<br />
isolated rat hepatocytes (Bonfanti et al., 1992).<br />
In vivo studies concerning its biotransformation in rat, mouse and<br />
monkey, however, pointed to the observation that Alachlor ® is metabolized<br />
via different pathways in rodents and monkeys, suggesting a lower risk for<br />
man than assumed (internal report Monsanto, 1988). In the future, this<br />
type <strong>of</strong> biotransformation study could easily be performed with short- and<br />
long-term cultures <strong>of</strong> hepatocytes derived from different species, including<br />
man, providing relevant human information without interspecies<br />
extrapolation.<br />
Long-term hepatocyte cultures<br />
V.ROGIERS ET AL. 209<br />
During culture, hepatocytes undergo phenotypic changes as a function <strong>of</strong><br />
culture time affecting selectively components <strong>of</strong> phase 1 and/or phase 2<br />
biotransformation (Nakamura et al., 1983; Guillouzo, 1986; Mooney et<br />
al., 1992; Kocarek et al., 1993; Rogiers and Vercruysse, 1993). These<br />
changes are interpreted as dedifferentiation. In the literature, data have
210 USE OF LONG-TERM CULTURES OF HEPATOCYTES IN TOXICITY TESTING<br />
been presented that part <strong>of</strong> this loss <strong>of</strong> functionality can be prevented by<br />
several factors including soluble medium factors, extracellular matrix<br />
components and cell-cell interactions (reviews Guillouzo et al., 1990;<br />
Rogiers, 1993). At present no ‘ideal’ long-term hepatocyte culture model<br />
exists but valuable alternatives are being developed which take into<br />
account some <strong>of</strong> the factors mentioned. A recent development consists <strong>of</strong><br />
hepatocytes cultured in a collagen gel sandwich configuration (Dunn et al.,<br />
1991; Lee et al., 1992). This promising system is claimed to maintain longterm<br />
differentiation probably due to the reinstatement <strong>of</strong> the cellular<br />
polarity <strong>of</strong> the hepatocytes as a function <strong>of</strong> the extracellular matrix (Dunn<br />
et al., 1991; Lee et al., 1992). To date, only few results concerning xenobiotic<br />
metabolism, are available (Koebe et al., 1994). Another recently introduced<br />
model substantially improved the maintenance <strong>of</strong> xenobiotic metabolism<br />
by culturing hepatocytes on a mixture <strong>of</strong> crude membrane fractions with<br />
collagen type 1, combined with the use <strong>of</strong> culture medium supplemented<br />
with aprotinin and selenium (Saad et al., 1993).<br />
Also co-cultures <strong>of</strong> hepatocytes with rat epithelial cells <strong>of</strong> primitive<br />
biliary origin represent a rather new and valuable tool in xenobiotic<br />
biotransformation research and testing (Bégué et al., 1984a). The model<br />
has been developed in order to mimic better the microenvironment <strong>of</strong> the<br />
liver cells in vivo. It is until now, the only long-term hepatocyte culture<br />
system <strong>of</strong> which enough biotransformation data exist. Co-cultures <strong>of</strong><br />
hepatocytes retain to a great extent the morphological and biochemical<br />
characteristics <strong>of</strong> adult hepatocytes in vivo, including phase 1 and phase 2<br />
xenobiotic metabolism pathways (Guillouzo, 1986; Rogiers et al., 1992;<br />
Akrawi et al., 1993a). The following text reviews the actual knowledge<br />
concerning xenobiotic biotransformation in cocultured hepatocytes with<br />
emphasis on the authors’ own research.<br />
Phase 1<br />
reactions in co-cultured hepatocytes<br />
It has been claimed that as much as 100 per cent <strong>of</strong> the CYP content and Naminopyrine<br />
demethylation activity can be maintained in co-cultured rat<br />
hepatocytes with primitive biliary duct cells (Bégué et al., 1984b). Some<br />
other phase 1 enzymatic activities also appear to be maintained since drugs<br />
such as ketotifen (Le Bigot et al., 1987) and testosterone (Utesch, 1992)<br />
were metabolized by several pathways. Maier (1988) however, has<br />
reported that aldrin epoxidase activity underwent a significant decrease as<br />
a function <strong>of</strong> culture time. Niemann et al. (1991) was able to show that, as<br />
was also the case for mono-cultured rat hepatocytes, enzymatic activities<br />
belonging to the 3-methyl-cholanthrene-inducible family were better<br />
maintained than those belonging to the phenobarbital inducible family. In<br />
our own experiments with adult rat hepatocytes co-cultured with rat liver
V.ROGIERS ET AL. 211<br />
epithelial cells (Rogiers et al., 1990b), it was found that a steady-state<br />
situation for at least 10 days was obtained in which the total CYP content<br />
and 7-ethoxycoumarin O-deethylase and aldrin epoxidase activities were<br />
maintained at 25,100 and 15 per cent, respectively, <strong>of</strong> their corresponding<br />
values in freshly isolated hepatocytes. Both the total CYP content and 7ethoxycoumarin<br />
O-deethylase activity, but not the aldrin epoxidase<br />
activity, were induced by exposure to phenobarbital (Rogiers et al., 1990b)<br />
or sodium valproate (Rogiers et al., 1988b). Furthermore, the combination<br />
<strong>of</strong> inducing agents with co-cultivation <strong>of</strong> rat hepatocytes with rat epithelial<br />
cell clones has recently been proposed as the best way for stabilization <strong>of</strong><br />
the CYP system. This method is better than the use <strong>of</strong> a perfusion system<br />
or changing the extracellular matrix from collagen to matrigel (Wegner et<br />
al., 1991). The results mentioned above indicate a degree <strong>of</strong> selection in the<br />
ability <strong>of</strong> co-cultured hepatocytes to maintain the expression and<br />
inducibility <strong>of</strong> individual members <strong>of</strong> the CYP superfamily. From the work<br />
<strong>of</strong> Akrawi (Akrawi et al., 1993a), using mRNA analysis <strong>of</strong> co-cultured rat<br />
hepatocytes, it appeared that the abundance <strong>of</strong> CYP2B mRNAs declined to<br />
about 30 per cent <strong>of</strong> the initial value by 4 days but that thereafter it remained<br />
constant. The inducibility by phenobarbital (Akrawi et al., 1993a) and<br />
sodium valproate (Rogiers et al., 1992) was also maintained. These results<br />
were confirmed by Western blotting (Akrawi et al. 1993b). RNase<br />
protection assays using probes capable <strong>of</strong> distinguishing between CYP2B1<br />
and CYP2B2 mRNAs demonstrated that the relative abundance and<br />
inducibility <strong>of</strong> each <strong>of</strong> the mRNAs were the same in co-cultures as in vivo.<br />
Co-cultured hepatocytes also maintained the expression <strong>of</strong> the CYP1A2<br />
gene and <strong>of</strong> genes coding for two other components <strong>of</strong> the CYP-mediated<br />
monooxygenase, namely NADPH cytochrome P450 reductase and<br />
cytochrome b 5 (Akrawi et al., 1993a). Both components were inducible by<br />
valproate and phenobarbital (Akrawi et al., 1993b; Shephard et al., 1994;<br />
Rogiers et al., 1994).<br />
In addition, we have shown using Western blotting (Akrawi et al.,<br />
1993b) and mRNA analysis (Akrawi et al., 1994; Shephard et al., 1994)<br />
that the expression <strong>of</strong> CYP4A and its specific inducibility (induced by<br />
valproate and not by phenobarbital) were maintained in co-cultured rat<br />
hepatocytes. By using antisense RNA probes that could discriminate<br />
between RNAs encoding different members <strong>of</strong> the CYP4A subfamily it was<br />
further demonstrated that CYP4A1, CYP4A2 and CYP4A3 were all<br />
induced by valproate, although to differing extents. None <strong>of</strong> these mRNAs<br />
was increased by phenobarbital (Akrawi et al., 1994; Shephard et al.<br />
1994). The results were very similar to those observed in vivo (Shephard et<br />
al., 1994).<br />
Flavin-containing monooxygenase (FMO), a less well-known phase 1<br />
biotransformation enzyme than the CYP system, is responsible for the<br />
oxygenation <strong>of</strong> drugs, pesticides and dietary components (Ziegler, 1980). It
212 USE OF LONG-TERM CULTURES OF HEPATOCYTES IN TOXICITY TESTING<br />
may activate as well as deactivate a number <strong>of</strong> important molecules such as<br />
thioether-containing pesticides (Hajjar and Hodgson, 1980), 3,3′-dichlorobenzidine<br />
(Iba and Thomas, 1988), N-methyl-4-aminobenzene (Kadlubar<br />
et al., 1976) and others. The expression <strong>of</strong> FMOs is also maintained better<br />
and for a longer time in co-cultures <strong>of</strong> rat hepatocytes than it is in monocultures<br />
<strong>of</strong> these cells (Coecke et al., 1993). In co-cultures a steady state<br />
situation is obtained at a level <strong>of</strong> approximately 40 per cent <strong>of</strong> its initial<br />
value in freshly isolated hepatocytes (Coecke et al., 1993).<br />
Hormonal regulation <strong>of</strong> FMO is retained (Coecke et al., 1995a,b). It was<br />
shown that 17β-oestradiol significantly decreased FMO activity in cocultures<br />
<strong>of</strong> male rat hepatocytes which was not the case for testosterone<br />
and 5α-dihydrotestosterone. These data are in accordance with in vivo<br />
results (Coecke et al., 1995b). In addition, the thyroid hormone thyroxine<br />
and its metabolite L-triiodothyronine were found to cause a significant<br />
decrease in FMO activity suggesting a suppressive role in the regulation <strong>of</strong><br />
FMO in rat liver. In vivo data on this subject are not available in the<br />
literature.<br />
Phase 2<br />
reactions in co-cultured hepatocytes<br />
The activity, expression and regulation <strong>of</strong> the different glutathione Stransferase<br />
(GST) isoenzymes in co-cultured rat hepatocytes, has been<br />
extensively studied by our group (Vandenberghe et al., 1988a,b, 1989,<br />
1990a,b, 1991). As far as the enzymatic activity is concerned, it was<br />
observed that the composition <strong>of</strong> the culture medium was <strong>of</strong> much less<br />
influence in co-cultures than was the case in mono-cultures (Vandenberghe<br />
et al., 1988b). In cocultures GST activity was maintained for a longer<br />
period and at a more stable level, comparable to the in vivo situation. This<br />
observation was confirmed later by Niemann et al. (1991) and Utesch and<br />
Oesch (1992), although the latter investigators reported a high variability<br />
in their results depending on the batch <strong>of</strong> epithelial cells. GST activity in cocultured<br />
rat hepatocytes was found to be increased or decreased by<br />
phenobarbital and valproate, respectively (Rogiers et al., 1988a; Rogiers et<br />
al., 1992).<br />
These results are in good accordance with previously obtained in vivo<br />
data (Rogiers et al., 1988a; Rogiers et al., 1992). They were also confirmed<br />
at the protein level using the Western blot technique (Rogiers et al., 1995).<br />
Furthermore, GST activity is increased significantly, in a dose dependent<br />
way by ethanol. Both GST protein and mRNA amounts (in particular, GST<br />
subunits 3 and 4) were increased by this compound (Coecke et al., 1995c).<br />
Results obtained using different substrates suggested that the GST isoenzyme<br />
pr<strong>of</strong>ile changes as soon as hepatocytes are seeded in culture<br />
(Vandenberghe et al., 1988b). By a combination <strong>of</strong> GSH agarose affinity
V.ROGIERS ET AL. 213<br />
chromatography and reversed phase HPLC, the GST subunits <strong>of</strong> cocultured<br />
rat hepatocytes were purified and separated (Vandenberghe et al,<br />
1988a). Alterations comparable to those observed for mono-cultures<br />
suggested changes towards a more ‘foetal-like’ state. Less variations,<br />
however, were noticed in the GST subunit pattern <strong>of</strong> co-cultured<br />
hepatocytes when various media conditions were compared. Incorporation<br />
<strong>of</strong> 35 S-methionine in the medium showed the ability <strong>of</strong> co-cultured rat<br />
hepatocytes to synthesize the different GST subunits and suggested that<br />
changes in GST subunit expression under various culture conditions were<br />
the result <strong>of</strong> in vitro ‘de novo’ synthesis (Vandenberghe et al., 1990a).<br />
Northern blot analysis, using specific cDNA probes showed that the<br />
mRNA levels encoding GST subunits 1/2, 3/4 and 7 were very dependent<br />
on the culture medium.<br />
Again in co-cultures, the changes observed were much less marked than<br />
was the case for mono-cultures (Morel et al., 1989; Vandenberghe et al.,<br />
1990b). As already mentioned for conventionally cultured rat hepatocytes,<br />
phenobarbital had inducing effects on all the GST subunits, but to a<br />
different extent for each subunit (Vandenberghe, 1989). The increased<br />
steady-state mRNA levels observed in co-cultures after phenobarbital<br />
exposure were the result <strong>of</strong> an increased transcriptional activity <strong>of</strong> the GST<br />
genes together with a stabilizing effect <strong>of</strong> the compound (Vandenberghe et<br />
al., 1991).<br />
Also <strong>of</strong> interest is that the hormonal regulation <strong>of</strong> GST is maintained in<br />
co-cultures <strong>of</strong> male rat hepatocytes. 17β-Oestradiol, triiodothyronine and<br />
thyroxine cause a significant decrease in GST activity. Both the overall GST<br />
activity and in particular that <strong>of</strong> GST 3–3 and 3–4 are decreased (Coecke<br />
et al., 1995c). In contrast, male sex hormones and human growth hormone<br />
had little effect on the overall activity. The effects <strong>of</strong> triiodothyronine and<br />
thyroxine were particularly oriented towards GST subunits 3 and 4 and<br />
towards an as yet unidentified GST subunit, which was significantly<br />
increased (Coecke et al., 1995c). 17β-Oestradiol shifted the GST subunit<br />
pattern towards the one observed in freshly isolated cells whereas growth<br />
hormone had no specific effect on the individual protein classes (Coecke et<br />
al., 1995c).<br />
These results clearly show a hormonal regulation <strong>of</strong> GST in co-cultured<br />
rat hepatocytes, although previous work with mono-cultures failed to<br />
prove any direct effect (Gebhardt et al., 1990). The effects are most<br />
pronounced for the Mu-class GSTs. In man, Mu-class GST genes are<br />
structurally very similar to the rat genes and are <strong>of</strong> particular interest<br />
because 45 per cent <strong>of</strong> the European population fails to express a<br />
transferase at the GST M 1 locus (Zhong et al., 1993). It is this class <strong>of</strong> GSTs<br />
that is very effective in deactivating mutagenic and carcinogenic epoxides.<br />
UDP glucuronyltransferases (UDP-GT) have been much less studied in<br />
cocultures than GST. From the work <strong>of</strong> Niemann et al. (1991) it appears
214 USE OF LONG-TERM CULTURES OF HEPATOCYTES IN TOXICITY TESTING<br />
that 1-naphthol UDP-GT activity is maintained well in co-cultures whereas<br />
this is not true for morphine UDP-GT, although for both a steady state<br />
situation was reached. These data point to a shift towards a more<br />
differentiated pattern since 1-naphthol is considered to be more specific for<br />
the late foetal form <strong>of</strong> UDP-GT and morphine for the neonatal form. The<br />
maintenance <strong>of</strong> 1-naphthol UDP-GT has been confirmed by Utesch and<br />
Oesch (1992). Other studies on preservation <strong>of</strong> phase 2 enzymatic activity<br />
in co-cultures have dealt with the identification <strong>of</strong> the metabolites formed<br />
when drugs are added to the culture medium. In human co-culture it could<br />
be shown that the glucuronide metabolite <strong>of</strong> ketotifen was still present<br />
after 3 weeks whereas it became undetectable after 6 days in mono-culture<br />
(Bégué et al., 1984b). Similar observations have been made for other drugs<br />
such as caffeine and theophylline (Ratanasavanh et al., 1990).<br />
Conclusions<br />
At present no ideal culture system for hepatocytes can be proposed. In all<br />
models reported in the literature, phenotypic changes occur, affecting the<br />
various components <strong>of</strong> phase 1 and phase 2 xenobiotic metabolism to a<br />
different extent. An interesting conclusion, however, remains from the<br />
observation that, when co-cultures and mono-cultures <strong>of</strong> hepatocytes are<br />
compared, cocultures exhibit higher biotransformation capacities which are<br />
better and preserved for longer than is the case for mono-cultures. The<br />
inducibility by common inducers is fairly well maintained and seems, to a<br />
certain extent, comparable with the in vivo situation. In addition, hormonal<br />
regulation <strong>of</strong> phase 1 and phase 2 key enzymes seems to be well maintained<br />
and comparable with the in vivo situation. Co-cultures <strong>of</strong> hepatocytes with<br />
rat liver epithelial cells are therefore already <strong>of</strong> importance as an alternative<br />
model for risk assessment. In particular, when long-term effects <strong>of</strong> a<br />
chemical are to be expected.<br />
Some experience already exists concerning the application <strong>of</strong> co-cultured<br />
hepatocytes for the study <strong>of</strong> pharmaceuticals. As far as chemical products<br />
other than pharmaceuticals are concerned, experience is lacking although<br />
interesting results are to be expected particularly in those cases where<br />
chemicals can interfere with the human organism via the food chain or by<br />
occupational exposure. In vitro exploration <strong>of</strong> this new field in toxicology<br />
is a challenge for the coming years.<br />
Notes<br />
1a. 88/379/EEC<br />
Directive du Conseil du 7 juin 1988 concernant le rapprochement des<br />
dispositions législatives, réglementaires et administratives des Etats membres
elatives à la classification, à l’emballage et à 1’étiquetage des préparations<br />
dangereuses. Journal Officiel des Communautés européennes no L187, 16<br />
juillet 1988, p. 14.<br />
1b. 93/18/EEC<br />
Directive 93/18/CEE de la Commission du 5 avril 1993 portant troisième<br />
adaptation au progrès technique de la directive 88/379/CEE du Conseil<br />
concernant le rapprochement des dispositions législatives, réglementaires et<br />
administratives des Etats membres relatives à la classification, à l’emballage<br />
et à 1'étiquetage des preparations dangereuses. Journal Officiel des<br />
Communautés européennes no L104, 29 Avril 1993, p. 46.<br />
2a. 91/414/EEC<br />
Richtlijn van de Raad van 15 juli 1991 betreffende het op de markt<br />
brengen van gewasbeschermingsmiddelen. Publikatieblad van de Europese<br />
Gemeenschappen no L230, 19 Augustus 1991, p. 1.<br />
2b. 93/71/EEC<br />
Directive 93/71/CEE de la Commission du 27 juillet 1993 modifiant la<br />
directive 91/414/CEE du Conseil concernant la mise sur le marché<br />
des produits phytopharmaceutiques. Journal Officiel des Communautés<br />
européennes no L221, 31 Août 1993, p. 27.<br />
3. 93/C239/03<br />
Voorstel voor een richtlijn van de Raad betreffende het op de markt<br />
brengen van biociden. Publicatieblad van de Europese Gemeenschappen no<br />
C239, 3 September 1993, p. 3.<br />
4a. 76/768/EEC<br />
Richtlijn van de Raad van 27 juli 1976 betreffende de onderlinge<br />
aanpassing van de wetgevingen der Lid-Staten inzake kosmetische produkten.<br />
Publikatieblad van de Europese Gemeenschappen no L262, 27 juli 1976, p.<br />
169.<br />
4b. 93/35/EEC<br />
Directive 93/35/CEE du Conseil du 14 juin 1993 modifiant pour la sixième<br />
fois la directive 76/768/CEE concernant le rapprochement des lé-gislations des<br />
Etats membres relatives aux produits cosmétiques. Journal Officiel des<br />
Communautés européennes no L151, 23 juin 1993 p. 32.<br />
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Press.
PART FIVE<br />
Mechanisms <strong>of</strong> toxicity <strong>of</strong> industrial<br />
chemicals
17<br />
Peroxisome Proliferation<br />
BRIAN G.LAKE and ROGER J.PRICE<br />
BIBRA International, Carshalton, Surrey<br />
Introduction<br />
Peroxisomes (sometimes referred to as ‘microbodies’) are single<br />
membranelimited cytoplasmic organelles which are characterised by their<br />
content <strong>of</strong> catalase and a number <strong>of</strong> hydrogen peroxide generating oxidase<br />
enzymes (Cohen and Grasso, 1981; Reddy and Lalwani, 1983). In rat liver,<br />
peroxisomes are normally spherical or oval in shape, approximately 0.5 µm<br />
in diameter and contain a finely granular matrix with a crystalline nucleoid<br />
core. A number <strong>of</strong> reviews have been published dealing with various<br />
aspects <strong>of</strong> hepatic peroxisome proliferation (Cohen and Grasso, 1981;<br />
Reddy and Lalwani, 1983; Hawkins et al., 1987; Stott, 1988; Lock et al.,<br />
1989; Moody et al., 1991; Bentley et al., 1993; Lake, 1993). This chapter<br />
will focus on mechanisms <strong>of</strong> hepatocarcinogenesis, species differences in<br />
response and risk assessment <strong>of</strong> rodent peroxisome proliferators.<br />
Peroxisome proliferation in rodent liver<br />
Since the initial observations on the hepatic effects <strong>of</strong> the hypolipidaemic<br />
agent cl<strong>of</strong>ibrate (Paget, 1963; Hess et al., 1965) many compounds have<br />
been shown to produce hepatic peroxisome proliferation in rats and mice.<br />
Liver enlargement is due to both hyperplasia and hypertrophy and<br />
organelle proliferation is associated with a differential induction <strong>of</strong><br />
peroxisomal enzyme activities. Peroxisomes, like mitochondria, contain a<br />
complete fatty acid β-oxidation cycle (Lazarow and DeDuve, 1976). While<br />
the enzymes <strong>of</strong> the β-oxidation cycle (normally assessed as cyanideinsensitive<br />
palmitoyl-CoA oxidation) are markedly induced, only small<br />
changes are observed in other peroxisomal enzyme activities such as<br />
catalase and D-amino acid oxidase. Apart from stimulating peroxisomal<br />
fatty acid metabolism, peroxisome proliferators also increase microsomal<br />
fatty acid ( -l)- and particularly -hydroxylase activities. This is due to<br />
induction <strong>of</strong> cytochrome P-450 isoenzymes in the CYP4A subfamily and is
224 PEROXISOME PROLIFERATION<br />
normally measured as lauric acid 12-hydroxylase (Sharma et al., 1988a, b;<br />
Gibson, 1989). Peroxisome proliferators also markedly increase carnitine<br />
acetyltransferase activity which is localised in peroxisomal, mitochondrial<br />
and microsomal fractions (Ishii et al., 1980; Bieber et al., 1981). In rat liver<br />
good correlations have been reported between the induction <strong>of</strong><br />
peroxisomal fatty acid β-oxidation and organelle proliferation and between<br />
the induction <strong>of</strong> peroxisomal and microsomal fatty acid oxidising enzyme<br />
activities (Lake et al., 1984a; Lin, 1987; Sharma et al., 1988a, b; Dirven et<br />
al., 1992).<br />
Several laboratories have demonstrated that the characteristics <strong>of</strong><br />
peroxisome proliferation in vivo may also be observed in vitro in primary<br />
rat and mouse hepatocyte cultures. Indeed, hepatocyte cultures have been<br />
employed for studying various aspects <strong>of</strong> peroxisome proliferation<br />
including structureactivity relationships and species differences in response<br />
(Gray et al., 1982; Elcombe, 1985; Bieri, 1993; Lake and Lewis, 1993;<br />
Foxworthy and Eacho, 1994).<br />
Rodent peroxisome proliferators<br />
Many different classes <strong>of</strong> chemicals have been found to produce<br />
peroxisome proliferation in the rat and mouse (Cohen and Grasso, 1981;<br />
Reddy and Lalwani, 1983; Stott, 1988; Moody et al., 1991; Bentley et al.,<br />
1993; Lake and Lewis, 1993). Classes <strong>of</strong> industrial chemicals include<br />
plasticisers, chlorinated solvents (e.g. trichloroethylene, perchloroethylene),<br />
chlorinated paraffins and other chemicals (e.g. perfluoro-n-octanoic acid).<br />
Types <strong>of</strong> plasticisers known to produce peroxisome proliferation include<br />
phthalate esters (e.g. di-(2-ethyl-hexyl)phthalate (DEHP), di-(isodecyl)<br />
phthalate), adipate esters (e.g. di-(2-ethyl-hexyl)adipate (DEHA)) and other<br />
compounds (e.g. tri-(2-ethylhexyl) trimellitate). Apart from industrial<br />
chemicals other known rodent hepatic peroxisome proliferators include<br />
herbicides, hypolipidaemic and other categories <strong>of</strong> therapeutic agents,<br />
certain steroids, food flavours and natural products.<br />
While peroxisome proliferators appear to be structurally diverse, at least<br />
for some compounds, similarities in their three-dimensional structures have<br />
been reported (Lake et al., 1988; Lake and Lewis, 1993). Many studies<br />
have demonstrated structure-activity relationships for various classes <strong>of</strong><br />
peroxisome proliferators including industrial chemicals (Lake and Lewis,<br />
1993). A characteristic feature <strong>of</strong> many, but not all, peroxisome<br />
proliferators is the presence <strong>of</strong> an acidic function (Lake et al., 1988; Lock<br />
et al., 1989). This acidic function is normally a carboxyl group, either<br />
present as a free carboxyl group in the parent structure or one that is<br />
unmasked by metabolism. Alternatively, the chemical may contain a<br />
chemical grouping which is a bioisostere <strong>of</strong> a carboxyl group (Thornber,
1979), such as tetrazole or a sulphonamide moiety (Eacho et al., 1986;<br />
Lock et al., 1989).<br />
It should be noted that rodent liver peroxisome proliferators exhibit<br />
marked compound potency differences. While potent peroxisome<br />
proliferators include compounds developed as hypolipidaemic agents (e.g.<br />
cipr<strong>of</strong>ibrate, Wy-14,643), plasticisers such as DEHP are less potent and<br />
chemicals such as acetylsalicylic acid are even less potent (Reddy et al.,<br />
1986; Barber et al., 1987; Lake and Lewis, 1993). For example, in a 30day<br />
feeding study, a similar magnitude <strong>of</strong> induction <strong>of</strong> palmitoyl-CoA<br />
oxidation was observed in rats fed 0.001 per cent cipr<strong>of</strong>ibrate and 0.5 per<br />
cent DEHP diets, whereas for DEHA a dietary level <strong>of</strong> >1.0 per cent but
226 PEROXISOME PROLIFERATION<br />
demonstrated to be effective in rat liver tumour promotion studies (Cattley<br />
and Popp, 1989; Bentley et al., 1993; Popp and Cattley, 1993).<br />
Mechanisms <strong>of</strong> hepatocarcinogenesis<br />
Several hypotheses have been proposed to account for why peroxisome<br />
proliferators can produce liver tumours in rodents. These mechanisms<br />
include:<br />
(a) Induction <strong>of</strong> sustained oxidative stress to hepatocytes (Reddy and<br />
Lalwani, 1983; Reddy and Rao, 1989).<br />
(b) A role <strong>of</strong> increased cell proliferation (Marsman et al., 1988; Popp and<br />
Marsman, 1991).<br />
(c) The promotion <strong>of</strong> spontaneously formed preneoplastic liver lesions<br />
(Schulte-Hermann et al., 1989; Cattley et al., 1991; Grasl-Kraupp et<br />
al., 1993).<br />
(d) A combination <strong>of</strong> two or all <strong>of</strong> the above factors.<br />
The oxidative stress hypothesis is based on the observation that the chronic<br />
administration <strong>of</strong> peroxisome proliferators produces a sustained oxidative<br />
stress in rodent hepatocytes due to an imbalance in the production and<br />
degradation <strong>of</strong> hydrogen peroxide (Reddy and Lalwani, 1983; Reddy and<br />
Rao, 1989). Peroxisome proliferators markedly induce the peroxisomal<br />
fatty acid β-oxidation cycle, but produce only a small increase in catalase<br />
activity. The first enzyme <strong>of</strong> the β-oxidation cycle, acyl-CoA oxidase,<br />
produces hydrogen peroxide and hence the cyclic oxidation <strong>of</strong> a single fatty<br />
acid molecule can result in the production <strong>of</strong> several molecules <strong>of</strong> hydrogen<br />
peroxide (Lazarow and DeDuve, 1976). Any excess hydrogen peroxide not<br />
destroyed by peroxisomal catalase can diffuse through the peroxisomal<br />
membrane into the cytosol where it will be a substrate for cytosolic<br />
selenium-dependent glutathione peroxidase. However, this enzyme activity<br />
and that <strong>of</strong> other enzymes including superoxide dismutase and glutathione<br />
S-transferases are <strong>of</strong>ten reduced by the administration <strong>of</strong> peroxisome<br />
proliferators to rodents (Reddy and Rao, 1989; Bentley et al., 1993; Lake,<br />
1993). These enzyme changes are postulated to result in increased<br />
intracellular levels <strong>of</strong> hydrogen peroxide which, either directly or via<br />
reactive oxygen species (e.g. hydroxyl radical), can attack membranes and<br />
DNA (Reddy and Lalwani, 1983; Reddy and Rao, 1989).<br />
A number <strong>of</strong> experimental observations have provided support for the<br />
involvement <strong>of</strong> oxidative stress in the hepatotoxicity <strong>of</strong> peroxisome<br />
proliferators (Reddy and Rao, 1989; Lake, 1993). For example,<br />
peroxisome proliferators have been reported in some studies to increase<br />
hepatic lipid peroxidation and lip<strong>of</strong>uscin deposition, to modulate levels <strong>of</strong><br />
hepatic antioxidants and to increase levels <strong>of</strong> 8-hydroxydeoxyguanosine in
B.G.LAKE AND R.J.PRICE 227<br />
hepatic DNA (Reddy and Rao, 1989; Bentley et al., 1993; Lake, 1993).<br />
However, the available data suggest that sustained oxidative stress is<br />
unlikely to be solely responsible for per oxisome proliferator-induced<br />
hepatocarcinogenesis in rodents. Although evidence <strong>of</strong> oxidative damage to<br />
hepatocytes has been observed in some studies, the magnitude <strong>of</strong> such<br />
effects does not correlate with the potency <strong>of</strong> the compound to produce<br />
tumours. For example, oxygen radical attack on DNA is known to result in<br />
a variety <strong>of</strong> modified DNA bases including 8-hydroxydeoxyguanosine.<br />
However, at bioassay dose levels both DEHP and DEHA produce similar<br />
increases in hepatic 8-hydroxydeoxyguanosine levels, but only DEHP<br />
produced liver tumours in male F344 rats (NTP, 1982a, b; Takagi et al.,<br />
1990; Lake, 1993).<br />
Many studies have demonstrated that cell proliferation is an important<br />
factor in the development <strong>of</strong> tumours by both genotoxic and nongenotoxic<br />
agents (Cohen and Ellwein, 1990, 1991). For example, an enhanced rate <strong>of</strong><br />
cell replication can increase the frequency <strong>of</strong> spontaneous lesions and the<br />
probability <strong>of</strong> converting DNA adducts from both endogenous and<br />
exogenous sources into mutations before they can be repaired (Cohen and<br />
Ellwein, 1990, 1991; Popp and Marsman, 1991). Peroxisome proliferators<br />
are known to produce a burst <strong>of</strong> cell replication in rodent hepatocytes<br />
during the first few days <strong>of</strong> administration (Reddy and Lalwani, 1983;<br />
Eacho et al., 1991). In some studies peroxisome proliferators have also<br />
been shown to produce a sustained stimulation <strong>of</strong> replicative DNA<br />
synthesis (Lake, 1993). Apart from intrinsic compound potency, dose is an<br />
important factor in determining whether a particular compound can<br />
produce either a transient or a sustained stimulation <strong>of</strong> replicative DNA<br />
synthesis in rodent hepatocytes. For example, low doses <strong>of</strong> nafenopin and<br />
Wy-14,643 do not produce a sustained stimulation <strong>of</strong> cell replication,<br />
whereas higher doses do produce this effect (Eacho et al., 1991; Price et al.,<br />
1992; Wada et al., 1992; Lake et al., 1993).<br />
Several studies have demonstrated the presence <strong>of</strong> numerous foci <strong>of</strong><br />
putative preneoplastic cells in the livers <strong>of</strong> untreated old rats and mice<br />
(Schulte-Hermann et al., 1983; Grasl-Kraupp et al., 1993). These lesions<br />
are considered to represent spontaneously initiated cells as they have<br />
similar biological characteristics to those <strong>of</strong> cells initiated by genotoxic<br />
carcinogens (Grasl-Kraupp et al., 1993). The ability <strong>of</strong> peroxisome<br />
proliferators to produce tumours in young compared to old rats has been<br />
investigated in studies with nafenopin (Kraupp-Grasl et al., 1991) and<br />
Wy-14,643 (Cattley et al., 1991). In both studies more adenomas and<br />
carcinomas were produced in old as against young rats.
228 PEROXISOME PROLIFERATION<br />
Species differences in response<br />
Many studies have examined species differences in hepatic peroxisome<br />
proliferation (Cohen and Grasso, 1981; Rodricks and Turnbull, 1987;<br />
Stott, 1988; Lock et al., 1989; Moody et al., 1991; Bentley et al., 1993).<br />
Based on both marker enzyme activities and ultrastructural examination<br />
the rat and mouse are clearly responsive species, the Syrian hamster<br />
appears to exhibit an intermediate response, whereas in most studies the<br />
guinea pig is either nonresponsive or refractory. For example, DEHP<br />
readily produces peroxisome proliferation in the rat and mouse, to a lesser<br />
extent in the Syrian hamster but not in the guinea pig (Osumi and<br />
Hashimoto, 1978; Lake et al., 1984b). Similar results have been obtained<br />
with more potent compounds including cipr<strong>of</strong>ibrate, clobuzarit, LY<br />
171883 and nafenopin (Orton et al., 1984; Eacho et al., 1986; Lake et al.,<br />
1989; Makowska et al., 1992).<br />
When assessing species differences in response a number <strong>of</strong> factors<br />
should be considered. These include the metabolism, disposition and dose<br />
<strong>of</strong> the test compound, sex differences, as well as intrahepatic differences in<br />
response. The importance <strong>of</strong> metabolism is illustrated by the industrial<br />
solvent trichloroethylene which produces peroxisome proliferation and<br />
liver tumours in the mouse but not in the rat (NCI, 1976; Elcombe, 1985).<br />
Metabolic studies demonstrated that the trichloroethylene was extensively<br />
metabolised to trichloroacetic acid in the mouse, whereas this was a minor<br />
saturable route <strong>of</strong> metabolism in the rat. That the difference in<br />
trichloroacetic acid formation was responsible for the observed species<br />
difference was demonstrated by the fact that this compound produced<br />
peroxisome proliferation in rat and mouse hepatocytes both in vivo and in<br />
vitro (Elcombe, 1985). An example <strong>of</strong> compound disposition is provided<br />
by DEHP which is known to be more extensively absorbed after oral<br />
administration in the rat than in the marmoset (Rhodes et al., 1986).<br />
However, the observed in vivo species differences in response are supported<br />
by the observation that metabolites <strong>of</strong> DEHP which produce peroxisome<br />
proliferation in rat hepatocytes in vitro have no significant effect in<br />
cultured marmoset hepatocytes (Elcombe and Mitchell, 1986). Generally,<br />
in vitro studies with primary hepatocyte cultures from the rat, mouse,<br />
Syrian hamster, guinea pig and marmoset have supported the results <strong>of</strong> in<br />
vivo studies in these species (Elcombe 1985; Elcombe and Mitchell, 1986;<br />
Lake et al., 1986; Bieri, 1993; Bentley et al., 1993; Foxworthy and Eacho,<br />
1994).<br />
Several studies have examined the ability <strong>of</strong> rodent peroxisome<br />
proliferators to produce effects in primates and humans. With respect to<br />
primates, studies with a number <strong>of</strong> compounds in both New (e.g.<br />
marmoset) and Old (e.g. Rhesus monkey) World monkeys have failed to<br />
provide any evidence <strong>of</strong> significant hepatic peroxisome proliferation
(Rodricks and Turnbull, 1987; Bentley et al., 1993). However, albeit at<br />
high doses two compounds, namely cipr<strong>of</strong>ibrate (Reddy et al., 1984) and<br />
DL-040 (Lalwani et al., 1985), have been reported to produce hepatic<br />
peroxisome proliferation in Cynomolgus and/or Rhesus monkeys. In<br />
humans, studies have been conducted in patients treated with several<br />
hypolipidaemic agents (all being rodent peroxisome proliferators) including<br />
cipr<strong>of</strong>ibrate, cl<strong>of</strong>ibrate, fen<strong>of</strong>ibrate and gemfibrozil (Bentley et al., 1993).<br />
While most studies have failed to detect any significant changes, cl<strong>of</strong>ibrate<br />
was reported to produce a small increase in the number <strong>of</strong> peroxisomes<br />
(Hanefeld et al., 1983) and cipr<strong>of</strong>ibrate to produce a small increase in the<br />
pro portion <strong>of</strong> the hepatocyte cytoplasm occupied by peroxisomes (cited in<br />
Bentley et al., 1993). However, owing to the large interindividual variation<br />
in peroxisome morphometrics observed in these studies, together with cell<br />
to cell variations and lobular variations, it is difficult to attach any clear<br />
biological significance to these findings (Bentley et al., 1993). Generally,<br />
peroxisome proliferators have not been reported to produce any significant<br />
effects on marker enzyme activities and/or peroxisomes in cultured primate<br />
and human hepatocytes (Bieri, 1993; Bentley et al., 1993; Foxworthy and<br />
Eacho, 1994).<br />
Some studies have also examined species differences in effects on cell<br />
replication. Both nafenopin and Wy-14,643 have been reported to<br />
stimulate replicative DNA synthesis in rat, but not in Syrian hamster,<br />
hepatocytes (Price et al., 1992; Lake et al., 1993). Although peroxisome<br />
proliferators can stimulate DNA synthesis in cultured rat hepatocytes,<br />
methylcl<strong>of</strong>enapate was reported to be ineffective in guinea pig, marmoset<br />
and human hepatocytes (Elcombe and Styles, 1989). Similarly, nafenopin<br />
has also been reported not to induce replicative DNA synthesis in human<br />
hepatocytes (Parzefall et al., 1991).<br />
Risk assessment <strong>of</strong> rodent liver peroxisome proliferators<br />
The key issues concerning the risk assessment <strong>of</strong> rodent liver peroxisome<br />
proliferators include:<br />
(a) Genotoxicity.<br />
(b) Likely human exposure.<br />
(c) Compound potency and no effect levels.<br />
(d) Precise mechanism(s) <strong>of</strong> liver tumour formation.<br />
(e) Species differences in response.<br />
B.G.LAKE AND R.J.PRICE 229<br />
Generally, peroxisome proliferators are considered to be non-genotoxic<br />
agents (Bentley et al., 1993; Budroe and Williams, 1993) and hence should<br />
be assessed differently from genotoxic carcinogens (Weisburger, 1994).<br />
Human exposure to rodent peroxisome proliferators depends on the
230 PEROXISOME PROLIFERATION<br />
intended usage <strong>of</strong> the particular compound. While hypolipidaemic agents<br />
are only administered to a restricted population <strong>of</strong> humans, exposure to<br />
industrial chemicals such as plasticisers is obviously far more widespread.<br />
For example, based on food surveillance surveys the daily human exposure<br />
to DEHA was reported to be 16 and 8.2 mg per person per day in 1987<br />
and 1990, respectively (MAFF 1987, 1990). In another study, where<br />
DEHA intake was assessed by measuring urinary levels <strong>of</strong> the major<br />
metabolite 2-ethylhexanoic acid, a median value <strong>of</strong> 2.7 mg per person per<br />
day was reported (L<strong>of</strong>tus et al., 1994).<br />
Apart from likely human exposure, consideration should be made <strong>of</strong> the<br />
relative potency <strong>of</strong> the particular compound to produce peroxisome<br />
proliferation and liver tumours in rodents. Plasticisers such as DEHP and<br />
DEHA are far less potent than certain therapeutic agents and<br />
experimentally used compounds (Reddy et al., 1986; Barber et al., 1987;<br />
Bentley et al., 1993; Lake and Lewis, 1993). Moreover rodent liver<br />
peroxisome proliferators exhibit clear no effect levels for both peroxisome<br />
proliferation and for tumour formation. For example, in the rat no effect<br />
levels for liver tumour formation have been observed in studies with<br />
several compounds including bezafibrate, cl<strong>of</strong>ibrate, DEHA and DEHP<br />
(Hartig et al., 1982; NTP 1982a, b). In addition, the threshold for tumour<br />
formation in rodents is appreciably greater than the threshold for<br />
peroxisome proliferation (Hartig et al., 1982; Reddy et al., 1986; Bentley<br />
et al., 1993).<br />
Several mechanisms have been proposed to account for why peroxisome<br />
proliferators produce tumours in rodent liver. If these various hypotheses<br />
are combined then a role for increased cell replication in peroxisome<br />
proliferatorinduced hepatocarcinogenesis may be readily identified. For<br />
example, if hepatocytes are transformed by either oxidative stress-induced<br />
damage or by alternative mechanisms, such initiated cells may be promoted<br />
to liver tumours by enhanced cell replication. Certainly peroxisome<br />
proliferators are effective promoters <strong>of</strong> certain populations <strong>of</strong> initiated cells<br />
and recent studies suggest that peroxisome proliferators can influence rates<br />
<strong>of</strong> both cell replication and cell death in particular populations <strong>of</strong><br />
hepatocytes (Grasl-Kraupp et al., 1993; Popp and Cattley, 1993; Marsman<br />
and Popp, 1994).<br />
With respect to species differences, rats and mice are clearly responsive<br />
species, whereas the majority <strong>of</strong> both in vivo and in vitro studies suggest<br />
that primates including man are either essentially refractory or certainly<br />
much less responsive to rodent peroxisome proliferators. However while<br />
effects on peroxisome morphology and marker enzyme activities have been<br />
extensively studied, few investigations have examined species differences in<br />
peroxisome proliferator-induced cell replication and liver tumour<br />
formation. As enhanced cell replication appears to play a role in peroxisome<br />
proliferator-induced hepatocarcinogenesis in rats and mice, it would
appear to be an important biomarker for assessing species differences in<br />
response. Rodent peroxisome proliferators do not appear to stimulate<br />
replicative DNA synthesis in vivo in Syrian hamster hepatocytes and in<br />
vitro in human hepatocytes (Elcombe and Styles 1989; Parzefall et al.,<br />
1991; Price et al., 1992; Lake et al., 1993). With respect to tumour<br />
formation, nafenopin and Wy-14,643 (two potent peroxisome<br />
proliferators) were reported not to produce liver lesions in the Syrian<br />
hamster although both compounds produced liver nodules and<br />
hepatocellular carcinoma after 60 weeks in the rat (Lake et al., 1993).<br />
Similarly, cl<strong>of</strong>ibrate was reported not to increase liver weight or produce<br />
liver tumours in marmosets after 6.5 years treatment (Tucker and Orton,<br />
1993) and in an ongoing study cipr<strong>of</strong>ibrate was found not to produce any<br />
morphological changes in marmoset liver after 3 years administration<br />
(Graham et al., 1994).<br />
In conclusion, the present literature suggests that rodent peroxisome<br />
proliferators are non-genotoxic agents which should be assessed differently<br />
from genotoxic compounds for human hazard (Weisburger, 1994).<br />
Assessment <strong>of</strong> likely human exposure and compound potency are also<br />
important factors together with information on compound no effect levels<br />
and evidence <strong>of</strong> species differences in response. Rodent liver peroxisome<br />
proliferators as a class <strong>of</strong> chemicals thus do not appear to pose any serious<br />
hazard for man. However, it would be desirable to elucidate further the<br />
mechanism(s) <strong>of</strong> peroxisome proliferator-induced hepatocarcinogenesis in<br />
susceptible species (i.e. the rat and mouse). From such studies the most<br />
appropriate biomarkers <strong>of</strong> liver tumour formation could be identified and<br />
examined in studies <strong>of</strong> species differences possibly including in vitro studies<br />
with human hepatocytes. Finally, further carcinogenicity studies in partially<br />
responsive (e.g. Syrian hamster) and non-responsive (e.g. guinea pig)<br />
species would strengthen the conclusion that peroxisome proliferators do<br />
not constitute any significant hazard to man.<br />
Acknowledgement<br />
We thank the UK Ministry <strong>of</strong> Agriculture, Fisheries and Food for financial<br />
support <strong>of</strong> BIBRA studies on hepatic peroxisome proliferation.<br />
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MARSMAN, D.S., CATTLEY, R.C., CONWAY, J.G. and POPP, J.A., 1988,<br />
Relationship <strong>of</strong> hepatic peroxisome proliferation and replicative DNA<br />
synthesis to the hepatocarcinogenicity <strong>of</strong> the peroxisome proliferators di(2ethylhexyl)phthalate<br />
and [4-chloro-6-(2,3-xylidino)-2-pyrimidinylthio]acetic<br />
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MOODY, D.E., REDDY, J.K., LAKE, B.G., POPP, J.A. and REESE, D.H., 1991,<br />
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Testing for induction <strong>of</strong> DNA synthesis in human hepatocyte primary cultures<br />
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Peroxisomes: Biology and Importance in <strong>Toxicology</strong> and Medicine, pp. 653–<br />
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PRICE, R.J., EVANS, J.G. and LAKE B.G., 1992, Comparison <strong>of</strong> the effects <strong>of</strong><br />
nafenopin on hepatic peroxisome proliferation and replicative DNA synthesis<br />
in the rat and Syrian hamster, Food and Chemical <strong>Toxicology</strong> 30, 937–44.<br />
REDDY, J.K. and LALWANI, N.D., 1983, Carcinogenesis by hepatic peroxisome<br />
proliferators: evaluation <strong>of</strong> the risk <strong>of</strong> hypolipidemic drugs and industrial<br />
plasticisers to humans, CRC Critical Reviews in <strong>Toxicology</strong>, 12, 1–58.<br />
REDDY, J.K. and RAO, M.S., 1989, Oxidative DNA damage caused by persistent<br />
peroxisome proliferation: its role in hepatocarcinogenesis, Mutation Research,<br />
214, 63–8.<br />
REDDY, J.K., LALWANI, N.D., QURESHI, S.A., REDDY, M.K. and MOEHLE,<br />
C.M., 1984, Induction <strong>of</strong> hepatic peroxisome proliferation in non-rodent<br />
species, including primates, American Journal <strong>of</strong> Pathology, 114, 171–83.<br />
REDDY, J.K., REDDY, M.K., USMAN, M.I., LALWANI, N.D. and RAO, M. S.,<br />
1986, Comparison <strong>of</strong> hepatic peroxisome proliferative effect and its<br />
implication for hepatocarcinogenicity <strong>of</strong> phthalate esters, di(2-ethylhexyl)<br />
phthalate and di(2-ethylhexyl)adipate with a hypolipidemic drug,<br />
Environmental Health Perspectives, 65, 317–27.<br />
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JACKSON, S.J. and ELCOMBE, C.R., 1986, Comparative pharmacokinetics<br />
and subacute toxicity <strong>of</strong> di(2-ethylhexyl)phthalate (DEHP) in rats and<br />
marmosets: extrapolation <strong>of</strong> effects in rodents to man, Environmental Health<br />
Perspectives, 65, 299–308.<br />
RODRICKS, J.V. and TURNBULL, D., 1987, Interspecies differences in<br />
peroxisomes and peroxisome proliferation, <strong>Toxicology</strong> and <strong>Industrial</strong> Health,<br />
3, 197–212.<br />
SCHULTE-HERMANN, R., TIMMERMANN-TROSIENER, I. and<br />
SCHLUPPLER, J., 1983, Promotion <strong>of</strong> spontaneous preneoplastic cells in rat<br />
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SCHULTE-HERMANN, R., KRAUPP-GRASL, B., BURSCH, W., GERBRACHT,<br />
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SHARMA, R., LAKE, B.G., FOSTER, J. and GIBSON, G.G., 1988a, Microsomal<br />
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93.
18<br />
Neurotoxicity Testing <strong>of</strong> <strong>Industrial</strong> <strong>Compounds</strong>:<br />
in vivo Markers and Mechanisms <strong>of</strong> Action<br />
KORNELIS J.VAN DEN BERG, 1 JAN-BERT<br />
P.GRAMSBERGEN, 2 ELISABETH M.G.HOOGENDIJK, 1<br />
JAN H.C.M.LAMMERS, 1 WILLEM S.SLOOT 1 and<br />
BEVERLY M.KULIG 1<br />
1 TNO Nutrition and Food Research Institute, Rijswijk; 2<br />
Erasmus University, Rotterdam<br />
Introduction<br />
Neurotoxicity assessment is designed to provide an answer to the question<br />
<strong>of</strong> whether or not a particular chemical is able to evoke some form <strong>of</strong><br />
adverse effect specifically associated with the nervous system. The<br />
development <strong>of</strong> risk assessment procedures is a long-term goal in view <strong>of</strong><br />
the complexity <strong>of</strong> the target organ involved, e.g. the nervous system. As<br />
risk identification is a first step in the risk assessment paradigm <strong>of</strong><br />
chemicals (NAS, 1983), for neuro-toxicity this translates at present into<br />
procedures and methods aimed at characterization <strong>of</strong> altered<br />
neurobehaviour <strong>of</strong> exposed experimental animals and abnormalities in the<br />
morphology <strong>of</strong> the nervous system. It has been suggested that risk<br />
assessment procedures may take more the approach <strong>of</strong> ‘exposuredoseresponse’<br />
(Andersen, 1991) in which mechanisms play an important role at<br />
various levels, e.g. from disposition <strong>of</strong> the chemical through the body to<br />
target tissues, via biochemical interactions at the molecular level to a toxic<br />
response as altered behaviour. Understanding the mechanisms involved in<br />
this sequence <strong>of</strong> events is helpful to arrive in the future at quantitative risk<br />
assessment procedures, for instance biologically-based modelling (Borgh<strong>of</strong>f<br />
et al., 1991). In addition, a better knowledge <strong>of</strong> the mechanistic principles<br />
<strong>of</strong> neuro-toxic agents may lead to the development <strong>of</strong> specific biomarkers<br />
that would further improve the efficiency <strong>of</strong> procedures for neurotoxicity<br />
screening, given the magnitude <strong>of</strong> the number <strong>of</strong> potentially neurotoxic<br />
compounds (NRC, 1992).<br />
Only for a small select group <strong>of</strong> industrial chemicals are mechanisms <strong>of</strong><br />
neurotoxic action more or less characterized. Perhaps the oldest examples<br />
are the class <strong>of</strong> organophosphate (OP) pesticides where the molecular<br />
targets have been identified as acetylcholinesterase (AChE) and neuropathy<br />
target esterase (NTE), the latter being associated with organophosphate-
induced delayed neuropathy (OPIDN) (Cherniack, 1988). This has led to<br />
investigations into structure-activity relationships <strong>of</strong> OPs with NTE in vitro<br />
that have proved to be valuable in predicting OPIDN (Davis et al. 1985).<br />
Furthermore, certain tissue culture systems, such as neuroblastoma cell<br />
lines, contain NTE and AChE activity that may be suited for identification<br />
<strong>of</strong> OPs causing OPIDN (Veronesi, 1992). Biochemical assays for NTE and<br />
AChE, in conjunction with neurobehavioural observations and<br />
neuropathology, are currently incorporated into US and Japanese<br />
neurotoxicity testing guidelines.<br />
A second group <strong>of</strong> industrial chemicals for which indications exist on<br />
their mechanism <strong>of</strong> action include certain compounds causing peripheral<br />
neuro-pathy such as acrylamide, carbon disulphide and n-hexane. It is<br />
generally assumed that the primary action <strong>of</strong> these chemicals is the<br />
crosslinking <strong>of</strong> axonal proteins (Graham et al., 1982), a process that blocks<br />
axonal transport (Sickles, 1991) and may lead to degeneration <strong>of</strong> distal<br />
axons (Spencer and Schaumberg, 1980). The basic information thus<br />
collected is yet to be developed into a useful biomarker.<br />
The pyrethroid insecticides represent a third group <strong>of</strong> chemicals for<br />
which the neurotoxic mechanism <strong>of</strong> action has been elucidated. The major<br />
symptoms <strong>of</strong> pyrethroid intoxication, e.g. convulsions, tremors, paralysis,<br />
are primarily the result <strong>of</strong> interaction <strong>of</strong> the pyrethroids with sodiumchannels<br />
on nerve membranes (Lund and Narahashi, 1982). Whereas<br />
under normal circumstances a sodium channel is only opened during a<br />
depolarization event, pyrethroids prolong opening <strong>of</strong> these channels<br />
thereby causing repetitive excitation <strong>of</strong> nerve and nerve terminals. In tissue<br />
culture systems, e.g. neuroblastoma cell lines, direct electrophysiological<br />
studies on cells have been done to investigate the effectiveness <strong>of</strong> different<br />
pyrethroids for opening <strong>of</strong> sodium channels (Oortgiesen et al., 1989).<br />
Biomarkers <strong>of</strong> neurotoxicity<br />
It is clear that a mechanistic understanding <strong>of</strong> the neurotoxicity <strong>of</strong><br />
suspected chemicals is and will be proceeding at a slow pace. Eventual<br />
utilization <strong>of</strong> this knowledge in the form <strong>of</strong> biological markers for<br />
neurotoxicity risk assessment procedures is, necessarily, a long-term<br />
objective. In the mean time alternative approaches have been proposed<br />
recently that (a) may provide biomarkers that could aid to further define<br />
underlying mechanisms and (b) are directly linked to neurotoxic<br />
mechanisms <strong>of</strong> actions.<br />
Gliotypic and neurotypic proteins<br />
K.J.VAN DEN BERG ET AL. 239<br />
Insults to the brain by a large variety <strong>of</strong> agents or conditions, e.g. viral<br />
infection, auto-immune encephalitis, trauma and chemicals, evoke a fairly
240 NEUROTOXICITY TESTING OF INDUSTRIAL COMPOUNDS<br />
stereo typic response <strong>of</strong> astrocytes (Eng, 1988). The normally rather<br />
quiescent astrocytes change into reaction astrocytes with a characteristic<br />
morphological appearance, a process known as reactive gliosis or<br />
astrogliosis. It has been proposed that this fairly universal astrocytic<br />
reaction might be a useful parameter for risk identification purposes<br />
(O’Callaghan, 1992; Rosengren and Haglid, 1989). The hypertrophic<br />
response <strong>of</strong> astrocytes is biochemically characterized by a greatly enhanced<br />
synthesis <strong>of</strong> glial fibrillary acidic protein (GFAP), a major structural protein<br />
<strong>of</strong> intermediate filaments (Eng, 1988). Under a variety <strong>of</strong> experimentallyinduced<br />
insults to the central nervous system, increased GFAP<br />
immunoreactivity has consistently been found in association with neuronal<br />
damage (Eng, 1988; O’Callaghan and Miller, 1989). It must be kept in<br />
mind, however, that reactive astrogliosis represents an indirect indication<br />
<strong>of</strong> nerve cell damage or loss. A strategy has been proposed (Brock and<br />
O’Callaghan, 1987) to collect quantitative data on astrogliosis, for<br />
example for GFAP concentrations, in association with information on<br />
changes <strong>of</strong> neuron-specific proteins. e.g. synapsin I or synaptophysin. The<br />
general idea is that enhanced GFAP levels in combination with decreased<br />
synaptophysin concentrations in the brain would be strongly indicative <strong>of</strong> a<br />
neurotoxic event. Recent methodological developments have led to<br />
quantitative procedures for assessment <strong>of</strong> GFAP and synaptophysin in<br />
nerve tissues by dot-blot immunoassay and ELISA (Jahn et al. 1984;<br />
O’Callaghan, 1991).<br />
Cerebral calcium accumulation<br />
Nerve cells, like most animal cells, possess a complicated system<br />
comprising calcium gates, pumps and channels to maintain free cytosolic<br />
calcium ion levels within a low physiological range (Pounds, 1990). Under<br />
pathological conditions, including hypoxia-ischaemia and status<br />
epilepticus, calcium overload in vulnerable neurons has been observed that<br />
is thought to be associated with the process <strong>of</strong> cell death (both necrosis and<br />
apoptosis) (Boobis et al. 1989; Siesjo and Bengtsson, 1989). Cerebral<br />
calcium accumulation has, therefore, been proposed as a potential index <strong>of</strong><br />
brain pathology (Korf et al., 1986).<br />
Free radical formation<br />
Free radicals may be involved in mechanisms <strong>of</strong> toxicity, including neurotoxicity<br />
<strong>of</strong> chemicals and are gaining more attention as is obvious from a<br />
number <strong>of</strong> recent reviews on this topic (LeBel and Bondy, 1991; Halliwell<br />
et al. 1992; Aust et al., 1993). Basically, free radicals are molecules that<br />
contain one or more unpaired electrons, whereas most molecules are<br />
nonradicals. Because <strong>of</strong> the unpaired electron(s) free radicals are highly
eactive chemical intermediates. Once a free radical is formed it can initiate<br />
a chain <strong>of</strong> reactions as the free electron is passed from one molecule to the<br />
other. In vivo, free radicals involving oxygen species are continuously<br />
produced and evolution has also provided the body with defence<br />
mechanisms such as superoxide dismutase (SOD), catalase, and scavengers<br />
such as glutathione, ascorbate, etc.<br />
Oxidative stress by free radicals may be an important neuropathological<br />
mediator following exposure to a number <strong>of</strong> neurotoxic agents (LeBel and<br />
Bondy, 1991). Examples <strong>of</strong> neurotoxic chemicals suspected <strong>of</strong> causing free<br />
radicals include chlordecone, ethanol, methamphetamine, methyl mercury,<br />
toluene, triethyl lead and trimethyltin. Of special interest is a ‘designer’<br />
drug called MPTP 1 causing destruction <strong>of</strong> dopaminergic neurons <strong>of</strong> the<br />
basal ganglia with symptoms similar to Parkinson’s disease. It is the<br />
neurotoxic metabolite MPP +2 that has been found to induce cerebral<br />
oxygen radical formation in vitro. There is also suggestive evidence to<br />
indicate that free radical scavengers may provide protection <strong>of</strong> the basal<br />
ganglia against neuro-toxic effects <strong>of</strong> MPP + (LeBel and Bondy, 1991). The<br />
occurrence <strong>of</strong> these kinds <strong>of</strong> compounds has led, among other things, one<br />
to suspect possible involvement <strong>of</strong> environmental chemicals and factors<br />
associated with diet and/or lifestyle in Parkinson’s disease (Russell, 1992;<br />
Semchuk et al., 1993).<br />
Model neurotoxins<br />
In order to validate an approach based on changes in these proposed<br />
biomarkers experimental studies were performed using various model<br />
neuro-toxicants, including trimethyltin, kainic acid, heavy metals such as<br />
lead, methylmercury and manganese, and developmental neurotoxicants,<br />
e.g. polychlorinated biphenyls.<br />
Trimethyltin<br />
K.J.VAN DEN BERG ET AL. 241<br />
Trimethyltin (TMT) is known to cause in adult rats a rather selective<br />
neuronal degeneration in specific regions <strong>of</strong> the brain, notably in limbic<br />
structures such as the hippocampus where extensive loss <strong>of</strong> pyramidal cells<br />
in CA fields are observed by standard histopathological procedures<br />
(O’Callaghan, 1988). A single systemic dose <strong>of</strong> TMT (7.5 mg kg −1 ) given to<br />
adult rats caused, after a period <strong>of</strong> 3 weeks, an approximately three-fold<br />
enhanced level <strong>of</strong> GFAP in hippocampus (Figure 18.1, upper left panel). In<br />
a number <strong>of</strong> other brain regions, e.g. different parts <strong>of</strong> the cortex,<br />
thalamus, striatum, cerebellum and brain stem, no significant changes in<br />
GFAP were observed. Assessment <strong>of</strong> synaptophysin, a structural protein <strong>of</strong><br />
synaptic vesicles <strong>of</strong> neurons (Jahn et al. 1985) in the same brain regions is<br />
given in Figure 18.1 (lower left panel). TMT induced a significant
242 NEUROTOXICITY TESTING OF INDUSTRIAL COMPOUNDS<br />
Figure 18.1 Changes in cerebral glial fibrillary acidic protein (GFAP) and<br />
synaptophysin concentrations by trimethyltin (TMT) and kainic acid. Results are<br />
expressed as mean±SEM. Open bars, control; closed bars, 7.5 mg trimethyltin<br />
(TMT) per kg or 12 mg kainic acid per kg; HP, hippopcampus; AM, amygdala; *,<br />
p
Kainic acid<br />
K.J.VAN DEN BERG ET AL. 243<br />
Kainic acid (KA), a glutamate agonist, is a potent model neurotoxin that<br />
causes neurodegeneration in a number <strong>of</strong> limbic structures including<br />
hippocampus, amygdala, piriform cortex (Sperk et al. 1983; Ben-Ari,<br />
1985). It is generally assumed that the mechanism <strong>of</strong> action involves<br />
interaction <strong>of</strong> KA with a special class <strong>of</strong> glutamate receptors and release <strong>of</strong><br />
endogenous excitatory amino acids in levels that are detrimental to<br />
neurons (Meldrum and Garthwaite, 1990).<br />
Kainic acid did induce, in a single systemic dose (12 mg kg −1 ), highly<br />
increased concentrations <strong>of</strong> GFAP in a number <strong>of</strong> target structures. For<br />
instance in hippocampus and amygdala GFAP levels did rise to 650 and<br />
960 per cent <strong>of</strong> control levels (Figure 18.1, upper right panel). Recent<br />
results from this laboratory have revealed that in a time-course study (Van<br />
den Berg and Gramsbergen, 1993) maximum levels are obtained after 4<br />
weeks that remained highly elevated in most target brain regions for at<br />
least a period <strong>of</strong> 6 months (Figure 18.2). The quantitative data on GFAP<br />
concentrations were supported by increased GFAP immunoreactivity in<br />
hippocampal sections visualized by immunohistochemical procedures. In<br />
addition to the damage in the hippocampus, permanently enhanced GFAP<br />
levels were found in other brain regions, e.g. piriform cortex, septum<br />
(Gramsbergen and Van den Berg, 1994) known to be targets <strong>of</strong> KA<br />
neurotoxicity.<br />
Synaptophysin levels were significantly reduced by KA in hippocampus<br />
and amygdala to a comparable degree (Figure 18.1, lower right panel).<br />
These data indicate a decrease in synaptophysin content encountered in<br />
brain regions where neuronal elements are known to be lost by these<br />
model neurotoxins. The magnitude <strong>of</strong> the changes in synaptophysin<br />
concentrations were much smaller than those <strong>of</strong> GFAP in the same brain<br />
structures. This may be explained by the fairly selective neuronal loss in<br />
specific layers <strong>of</strong>, for instance, the hippocampus, by TMT and KA. The<br />
effect is thus rather diluted in a biochemical procedure. Once an effect is<br />
scored, more detailed biochemical analysis is possible by using a punch<br />
technique after microdissection <strong>of</strong> brain nuclei (Palkovits and Brownstein,<br />
1988). Alternatively, a follow-up by histopathological procedures, for<br />
example by the cupric silver degeneration stain, would provide further<br />
details <strong>of</strong> neuronal damage (O’Callaghan and Jensen, 1992). In the<br />
experiments described above there was no clear correlation between the<br />
graded regional GFAP response and decrease <strong>of</strong> synaptophysin<br />
concentration. This may suggest a differential region-specific response <strong>of</strong><br />
astrocytes towards neuronal injury.<br />
In our laboratory cerebral calcium accumulation was determined<br />
recently after a single systemic dose <strong>of</strong> KA (12 mg kg −1 , i.p.), given to adult<br />
rats. A rapid uptake <strong>of</strong> 45 Ca was observed in various regions <strong>of</strong> the brain. A
244 NEUROTOXICITY TESTING OF INDUSTRIAL COMPOUNDS<br />
Figure 18.2 Time-course <strong>of</strong> GFAP concentration in hippocampus by kainic acid. Rats<br />
were treated with a single systemic dose <strong>of</strong> kainic acid (12 mg kg −1 , i.p.). GFAP was<br />
determined by ELISA at the times indicated. Results are expressed as mean±SEM<br />
(from Van den Berg and Gramsbergen, 1993, with permission).<br />
time-course study, covering a period <strong>of</strong> 6 months, has indicated that in the<br />
hippocampus a peak <strong>of</strong> 45 Ca uptake occurred already after 4 days and<br />
normal values were reached only after another 2–4 months (Van den Berg<br />
and Gramsbergen, 1993). More recent results indicate that other limbic<br />
areas display similar kinetics, while 45 Ca uptake in striatum and cortex<br />
return faster to normal values, e.g. within 2 weeks. Only in the thalamus a<br />
long-term sustained 45 Ca uptake was present for a period <strong>of</strong> 6 months. In<br />
general a good correlation was found between the regions showing<br />
sustained 45 Ca accumulation and those known to be targets <strong>of</strong> KA<br />
neurotoxicity. When these data on 45 Ca accumulation were related to<br />
effects <strong>of</strong> KA on GFAP concentrations, also a good agreement was<br />
observed concerning the brain regions involved (Gramsbergen and Van den<br />
Berg, 1994). Histopathological examination <strong>of</strong> hippocampal sections<br />
revealed extensive neurodegeneration by KA in CA1, CA3 and CA4<br />
regions (Van den Berg and Gramsbergen, 1993).<br />
Heavy metals<br />
Lead may produce in children symptoms <strong>of</strong> acute encephalopathy after<br />
exposure to high doses and at lower dose levels learning disorders and
K.J.VAN DEN BERG ET AL. 245<br />
hyperactive behaviour. In adults neurotoxicity is more <strong>of</strong>ten encountered<br />
as peripheral neuropathy after chronic occupational exposure to lead<br />
(Marsh, 1985). A number <strong>of</strong> mechanisms have been implicated in lead<br />
neurotoxicity (Bressler and Goldstein, 1991), but as yet no central<br />
hypothesis has emerged.<br />
In order to investigate the effect <strong>of</strong> lead on CNS structural proteins, a<br />
subchronic dosing experiment with adult rats was performed in which<br />
animals received daily doses <strong>of</strong> lead acetate (4, 8, 12.5 mg kg −1 i.p.) for 28<br />
days. GFAP concentrations were subsequently determined in different brain<br />
regions. Already at the lowest dose level GFAP levels were found to be<br />
significantly increased in several brain regions, notably in different parts <strong>of</strong><br />
the cortex, hippocampus and striatum, while cerebellum and brain stem<br />
remained unaffected. Neurobehavioural assessment <strong>of</strong> animals, preceding<br />
the neurochemical analysis, also revealed significant alterations <strong>of</strong><br />
neuromuscular function, excitability and spontaneous activity.<br />
Methylmercury has caused a number <strong>of</strong> poisonings in man (Marsh, 1985)<br />
where it appears to affect in particular both the central and peripheral<br />
nervous system. In an animal experiment, adult rats were subchronically<br />
dosed with methylmercury (0.75 or 2 mg kg −1 ) for 28 days.<br />
Neurobehavioural assessment indicated that grip strength was significantly<br />
impaired. Neurochemical analysis <strong>of</strong> GFAP in the central nervous system<br />
was performed in selected brain regions and, in addition, in various<br />
segments <strong>of</strong> the spinal cord. Increased GFAP levels were observed in the<br />
cerebrum only in the frontal cortex, also in brain stem and in spinal cord. A<br />
further detailed analysis <strong>of</strong> brain stem sub-structures showed significantly<br />
enhanced GFAP levels in pons and medulla oblongata but not in midbrain.<br />
In spinal cord GFAP concentration was increased in specific<br />
sections, e.g. in the cervical and lumbar segments but not in the thoracic<br />
segment.<br />
The results with these particular examples <strong>of</strong> heavy metals have indicated<br />
the unsuspected presence <strong>of</strong> regions in the central nervous system with<br />
astrogliosis, as determined in a biochemical GFAP assay. The neuronal<br />
damage involved has not yet been confirmed independently, e.g. by<br />
assessment <strong>of</strong> synaptophysin. The possibilities remain, therefore, that<br />
reactive astrocytosis by these heavy metals may be indirectly a result <strong>of</strong><br />
breaching the integrity <strong>of</strong> the blood-brain barrier (Bressler and Goldstein,<br />
1991) or <strong>of</strong> a direct toxic action on astroglial cells (Selvin Testa et al.,<br />
1990; Stark et al., 1992).<br />
Manganese is a well recognized industrial neurotoxin associated with<br />
neurologic effects after prolonged exposure in occupational settings (Katz,<br />
1985). The clinical manifestations <strong>of</strong> manganism bear a large similarity to<br />
those <strong>of</strong> Parkinson’s disease (PD). The neurodegenerative disease PD is<br />
characterized by a selective loss <strong>of</strong> neurons in the basal ganglia.
246 NEUROTOXICITY TESTING OF INDUSTRIAL COMPOUNDS<br />
Experimental studies were done to investigate the vulnerability <strong>of</strong> the<br />
basal ganglia in manganism. Manganese (as Mn 2+ ) was applied<br />
intrastriatally to rats and region-specific brain damage was assessed by<br />
determining regional 45 Ca accumulation using quantitative<br />
autoradiographic procedures developed in this laboratory (Gramsbergen<br />
and Van der Sluijs-Gelling, 1993). It appeared that Mn 2+ induced a timedependent<br />
45 Ca accumulation in most <strong>of</strong> the regions constituting the basal<br />
ganglia, e.g. striatum and several other structures such as globus pallidus,<br />
entopeduncular nucleus, several thalamic nuclei and substantia nigra (Sloot<br />
et al., 1994).<br />
As monoaminergic neurotransmission plays a predominant role in the<br />
basal ganglia, recent studies have indicated that concentrations <strong>of</strong> biogenic<br />
amines such as dopamine and its major metabolites, serotonin and<br />
noradrenaline were also reduced in striatum by Mn 2+ . The kinetics <strong>of</strong> this<br />
process indicated that concentrations <strong>of</strong> most monoamines and metabolites<br />
were temporarily reduced except for dopamine and metabolites in striatum<br />
that remained at a permanently reduced level (>90 days) (Sloot et al.,<br />
1994). In order to demonstrate the specificity <strong>of</strong> the manganese effects,<br />
several other compounds were studied, including ferrous ions (Fe 2+ ). An<br />
equimolar dose <strong>of</strong> Fe 2+ , applied intrastriatally, produced, however, a much<br />
more extensive and widespread 45 Ca accumulation throughout the basal<br />
ganglia and, in addition, in nucleus accumbens and cerebral cortex.<br />
Ferrous ions were also three times more potent than manganese ions in<br />
causing depletion <strong>of</strong> dopamine in striatum (Sloot et al., 1994).<br />
The results based on both 45 Ca accumulation and biogenic amine levels<br />
are in concordance with the hypothesis that the basal ganglia, which are<br />
enriched in iron and iron-binding proteins, represent a selective target for<br />
manganese. The role and fate <strong>of</strong> endogenous iron in the brain, and the<br />
basal ganglia in particular, under toxic conditions including chronic<br />
manganese exposure, merits further investigations.<br />
Oxidative stress by free radical formation may play a role in these toxic<br />
events. Transition metals are known as strong promoters <strong>of</strong> reactive<br />
oxygen species. Especially iron, as the ferrous ion (Fe 2+ ), has been found to<br />
react with hydrogen peroxide to form the hydroxyl radical in the so-called<br />
Fenton reaction (Halliwell et al., 1992; Aust et al., 1993). It is thought that<br />
this mechanism plays an important direct role in iron poisoning (Aust et<br />
al., 1993). In an indirect way oxidative stress by iron may be initiated by<br />
chemicals that are able to ‘liberate’ iron from stores such as ferritin,<br />
transferrin, haemoglobin, etc. Recently, evidence has been obtained that a<br />
number <strong>of</strong> chemicals, including the pesticides paraquat and diquat, may<br />
release iron from ferritin in vivo as well as in vitro (Aust et al., 1993),<br />
involving organic radical and superoxide formation.<br />
These observations are particularly relevant for the interpretation <strong>of</strong> the<br />
observed neurotoxic effects <strong>of</strong> iron described above. Dopamine is relatively
easily subjected to a process <strong>of</strong> auto-oxidation. The decrease in dopamine<br />
levels by iron may have been caused by oxygen radicals. Circumstantial<br />
evidence suggests that a similar mechanism <strong>of</strong> action may underly<br />
manganese neurotoxicity. The decrease in dopamine in the basal ganglia by<br />
manganese is also thought to occur through catalysis <strong>of</strong> dopamine<br />
oxidation (LeBel and Bondy, 1991), possibly involving radical oxygen<br />
species. An open question is whether manganese might participate in the<br />
‘iron release’ hypothesis (Sloot and Gramsbergen, 1994). Current efforts<br />
are being made to determine free radical formation by iron and manganese<br />
and to relate this to dopamine depletion. For this purpose rats are given<br />
salicylic acid (SA) and subsequently the SA hydroxylation products are<br />
measured in cerebrospinal fluid and brain tissues as indices <strong>of</strong> hydroxyl<br />
radical formation.<br />
Developmental neurotoxins (PCBs)<br />
K.J.VAN DEN BERG ET AL. 247<br />
Concern has been raised about the long term consequences <strong>of</strong> low level<br />
intake <strong>of</strong> polychlorinated biphenyls (PCBs) with respect to neurotoxicity as<br />
it relates to nervous tissue development and intellectual performance in the<br />
juvenile and adult stages. Several epidemiological studies with infants have<br />
shown a negative correlation between PCB levels in cord blood and<br />
cognitive functions and a positive correlation between PCB levels and<br />
altered neurological parameters such as hypotonia and hyporeflexia (Rogan<br />
et al., 1988; Jacobson et al., 1990). Experimental studies in various species<br />
including primates have also provided arguments for neurotoxic effects in<br />
<strong>of</strong>fspring after perinatal exposure to PCBs (Tilson et al., 1990). In this<br />
laboratory evidence has recently been obtained to indicate dramatic<br />
reduction in sexual behaviour and reproduction in <strong>of</strong>fspring that was<br />
perinatally exposed to PCBs (Smits-van Proojie et al., 1993).<br />
In order to investigate eventual structural alterations in the CNS,<br />
pregnant Wistar WU rats were exposed to Aroclor 1254 on days 10–16 <strong>of</strong><br />
gestation. At a young age (3 weeks) and adult age (3 months) <strong>of</strong>fspring<br />
were sacrificed and various brain regions were dissected for assessment <strong>of</strong><br />
gliotypic and neurotypic proteins. In untreated control animals<br />
developmental aspects <strong>of</strong> astrocytes in the central nervous system were<br />
encountered. Both in hypothalamus and cerebellum <strong>of</strong> control animals<br />
GFAP levels were increased by 200–300 per cent between 3 weeks and 3<br />
months postnatally. A developmental GFAP increase was also found in<br />
brain stem, striatum and lateral olfactory tract, although to a lesser extent.<br />
In hippocampus and prefrontal cortex, GFAP levels remained virtually<br />
unchanged between 3 weeks and 3 months. These results, therefore,<br />
indicate that glial cell maturation and/or differentiation is not uniformly<br />
distributed over the whole brain. It appeared that glial cells at birth in rats<br />
were fully developed in brain regions dealing with cognitive functions, e.g.
248 NEUROTOXICITY TESTING OF INDUSTRIAL COMPOUNDS<br />
cortex and hippocampus, while in regions such as hypothalamus and<br />
cerebellum this process occurred entirely neonatally. It should be kept in<br />
mind that in the rat also neuronal maturation and differentiation are not<br />
completed at birth and continue to a large extent postnatally.<br />
Exposure <strong>of</strong> pregnant rats to Aroclor 1254 did lead to a number <strong>of</strong><br />
alterations in GFAP levels in various brain regions in <strong>of</strong>fspring as compared<br />
to control animals. The most striking differences were observed in brain<br />
stem. While in control animals <strong>of</strong> both sexes GFAP levels increased<br />
postnatally, the normal developmental increase in perinatally exposed<br />
animals was absent (Figure 18.3, upper panel). At 3 months the relative<br />
deficit in GFAP levels was 41 per cent for male and 30 per cent for female<br />
progeny. Already at the lowest dose <strong>of</strong> Aroclor 1254 (5 mg/kg) a maximum<br />
decrease was observed (Figure 18.3, upper panel). In addition to brain stem,<br />
a similar effect on GFAP levels was found in striatum although the relative<br />
deficit at 3 months was somewhat less. Quite an opposite pattern emerged<br />
in brain regions such as cerebellum, lateral olfactory tract and prefrontal<br />
cortex, where GFAP levels were increased relative to unexposed progeny.<br />
In hippocampus no significant changes in GFAP levels were encountered.<br />
The question arose whether the observed neurodevelopmental toxicity in<br />
rats by PCBs was specific for astroglial cells or also involved neuronal<br />
maturation, differentiation and death. For this purpose the same nervous<br />
tissues were used for quantitative assessment <strong>of</strong> a neuronal marker in the<br />
form <strong>of</strong> synaptophysin. In brains <strong>of</strong> untreated adult control rats, regionspecific<br />
differences were observed in synaptophysin concentrations, being<br />
high in the prefrontal cortex, striatum and hippocampus and relatively low<br />
in lateral olfactory tract, cerebellum and brain stem. Synaptogenesis for the<br />
brain as a whole largely takes place in the rat from birth until postnatal<br />
day 70 (Knaus et al., 1986). Regional differences in the speed <strong>of</strong> postnatal<br />
synaptogenesis from 3 weeks to 3 months were found in a number <strong>of</strong><br />
structures being most pronounced in cerebellum (190 per cent increase) and<br />
prefrontal cortex (170 per cent increase) while in brain stem little change was<br />
observed.<br />
As a result <strong>of</strong> perinatal exposure to Aroclor 1254, altered expression <strong>of</strong><br />
synaptophysin was observed. In most brain structures examined, including<br />
brainstem (Figure 18.3, lower panel) significant decreases in synaptophysin<br />
concentrations were found. This suggests that during development <strong>of</strong> the<br />
central nervous system, PCBs may interfere with the formation <strong>of</strong> synaptic<br />
vesicles, synaptogenesis, or formation <strong>of</strong> nerve terminals.<br />
A straightforward interpretation <strong>of</strong> the present results is not possible at<br />
this stage. However, what is clear is that perinatal exposure to PCBs may<br />
cause changes in the structural composition <strong>of</strong> the central nervous system<br />
both in the neuronal and the glial cell compartment. Apparently there are<br />
different effects on nerve cells <strong>of</strong> the CNS depending on the brain region<br />
involved. The brain stem and striatum are regions with decreased
K.J.VAN DEN BERG ET AL. 249<br />
Figure 18.3 Developmental effects on glial fibrillary acidic protein (GFAP) and<br />
synaptophysin concentrations in brain stem after perinatal exposure to Aroclor<br />
1254. Results are expressed as mean±SEM with the values <strong>of</strong> the control animals at<br />
day 21 set at 100 per cent. Open bars, control; shaded bars, 5 mg Aroclor 1254 per<br />
kg; closed bars, 25 mg Aroclor 1254 per kg; D 21 and D 90, postnatal days; *, p
250 NEUROTOXICITY TESTING OF INDUSTRIAL COMPOUNDS<br />
Table 18.1 Cerebral calcium accumulation<br />
a Systemic dose.<br />
b Intrastriatal dose.<br />
synaptophysin and increased GFAP in structures such as cerebellum, lateral<br />
olfactory tract, and prefrontal cortex. Such a situation could arise when<br />
neurons in these latter regions fail to receive proper input from developing<br />
neurons originating in brain stem and striatum. Further investigations into<br />
this interesting topic may provide further clues for the developmental<br />
toxicity <strong>of</strong> PCBs and related compounds that possibly could aid in the<br />
interpretation <strong>of</strong> neurobehavioural effects, for instance altered sexual<br />
behaviour and reproduction (Smits-van Prooije et al., 1993).<br />
The findings observed in cerebellum are consistent with a phase <strong>of</strong><br />
hypothyroidism during development but it is clear that a conclusive role<br />
for thyroid hormone remains to be further established. The present results<br />
furthermore suggest that alterations in synaptophysin/GFAP levels may be<br />
useful and sensitive parameters to study compounds suspected <strong>of</strong><br />
developmental neurotoxicity.<br />
Conclusion<br />
The results with various neurotoxins demonstrate that assessment <strong>of</strong><br />
gliotypic proteins such as GFAP may be a useful tool to identify and<br />
quantify persistent toxic insults <strong>of</strong> the CNS, especially when this is backed<br />
up by indications for loss <strong>of</strong> neuronal elements, e.g. decreased<br />
synaptophysin concentration. Also in circumstances <strong>of</strong> developmental<br />
neurotoxicity, caused by chemicals such as PCBs, an approach based on<br />
changes <strong>of</strong> gliotypic and neurotypic proteins may provide a promising<br />
biomarker. Of course, further investigations with various other compounds<br />
are required to substantiate these interesting findings.<br />
Cerebral calcium accumulation may be useful as an early indicator <strong>of</strong><br />
neurotoxicity on a more prospective basis as is suggested by various<br />
examples <strong>of</strong> neurotoxic compounds (summarized in Table 18.1). Because in<br />
unlesioned brain regions the background levels <strong>of</strong> calcium uptake remain
very low, areas where neurotoxic events take place are easily identified and<br />
quantified by autoradiographic or scintillation counting procedures.<br />
Finally, the role <strong>of</strong> oxidative stress by free radical formation in the<br />
nervous system merits further studies since it may open up possibilities for<br />
providing a biomarker that is linked to a mechanism <strong>of</strong> neurotoxic action.<br />
Acknowledgements<br />
The author is grateful for unpublished information provided by Dr Didema<br />
de Groot on neuropathology, by Dennis C.Morse, MSc and the students<br />
Wendelien Wesseling and Annemiek Plug, Agricultural University<br />
Wageningen, on developmental effects <strong>of</strong> PCBs; and for expert technical<br />
assistance by Alita van der Sluijs-Gelling.<br />
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SLOOT, W.N. and GRAMSBERGEN, J.B.P., 1994, Axonal transport <strong>of</strong><br />
manganese and its relevance to selective neurotoxicity in the rat basal ganglia,<br />
Brain Res., 657, 124–32.<br />
SLOOT, W.N., VAN DER SLUIJS-GELLING, A.J. and GRAMSBERGEN, J.B.P.,<br />
1994, Selective lesions by manganese and extensive damage by iron after<br />
injection into rat striatum or hippocampus, J. Neurochem., 62, 205–16.<br />
SMITS-VAN PROOIJE, LAMMERS, J.H.C.M., WAALKENS-BERENDSEN,<br />
D.H., KULIG, B.M. and SNOEIJ, N.J., 1993, Effects <strong>of</strong> the PCB 3,4,5,3′,4′,5′hexachlorobiphenyl<br />
on the reproduction capacity <strong>of</strong> Wistar rats,<br />
Chemosphere, 27, 395– 400.<br />
SPENCER, P.S. and SCHAUMBERG, H.H., 1980, Experimental and Clinical<br />
Neurotoxicology, Baltimore: Williams and Wilkins.<br />
SPERK, G., LASSMAN, H., BARAN, H., KISH, S.J., SEITELBERGER, F. and<br />
HORNYKIEWICZ, O., 1983, Kainic acid induced seizures: neurochemical and<br />
histopathological changes, Neuroscience, 10, 1301–15.<br />
STARK, M., WOLFF, J.E. and KORBMACHER, A., 1992, Modulation <strong>of</strong> glial cell<br />
differentiation by exposure to lead and cadmium, Neurotoxicol. Teratol, 14,<br />
247–52.<br />
TILSON, H.A., JACOBSON, J.L. and ROGAN, W.J., 1990, Polychlorinated<br />
biphenyls and the developing nervous system: cross-species comparisons,<br />
Neuro-toxicol. Teratol, 12, 239–48.<br />
VAN DEN BERG, K.J. and GRAMSBERGEN, J.B.P., 1993, Long-term changes in<br />
glial fibrillary acidic protein and calcium levels in rat hippocampus after a<br />
single systemic dose <strong>of</strong> kainic acid, Ann. N.Y. Acad. Sci., 679, 394–401.<br />
VERONESI, B., 1992, In vitro screening batteries for neurotoxicants,<br />
Neurotoxicology, 13, 185–96.
19<br />
Endocrine <strong>Toxicology</strong> <strong>of</strong> the Thyroid for<br />
<strong>Industrial</strong> <strong>Compounds</strong><br />
CHRISTOPHER K.ATTERWILL and SAMUEL<br />
P.AYLWARD<br />
CellTox Centre, University <strong>of</strong> Hertfordshire, Hatfield, Herts<br />
General introduction<br />
Classification <strong>of</strong> endocrine toxicity<br />
Xenobiotic-induced endocrine dysfunction and toxicity is a common<br />
finding in safety studies and an increasingly important consideration in the<br />
riskassessment process. The effective identification <strong>of</strong> potential endocrine<br />
toxicological effects depends upon xenobiotic effect classification in<br />
relation to normal endocrine function and pathology.<br />
Classifications <strong>of</strong> different types <strong>of</strong> endocrine toxicity have been<br />
proposed in previous publications on this subject. Capen and Martin<br />
(1989) proposed a detailed classification based on clinical endocrine<br />
function and pathology. On the other hand, Baylis and Tunbridge (1985)<br />
proposed a simpler classification based on the adverse endocrine reactions<br />
<strong>of</strong> xenobiotics which are observed clinically. From a toxicological or<br />
preclinical safety testing point <strong>of</strong> view, Atterwill and Flack (1993) favoured<br />
classifying endocrine toxicology <strong>of</strong> xenobiotics in a manner which is<br />
similar in concept to that for classifying other toxicological phenomena<br />
(CIOMS, 1983) taking into account the unique nature <strong>of</strong> the endocrine<br />
system. This is as follows (see also Figure 19.1):<br />
Class 1 Effects which can be predicted from the endocrine pharmacology<br />
<strong>of</strong> compounds. An example would be the oestrogens and progestogens<br />
which have a plethora <strong>of</strong> effects on metabolic parameters in addition to their<br />
actions on oestrogen sensitive target sites when administered at<br />
pharmacological doses (the therapeutic dose levels). Further examples are<br />
certain dopamine and 5-HT antagonists acting on hypothalamic and<br />
pituitary receptors which may disrupt ‘downstream’ endocrine functions.<br />
Class 2 Effects which again can be predicted from the endocrine<br />
pharmacology <strong>of</strong> the compound when administered at doses well in excess<br />
<strong>of</strong> the therapeutic dose level. An example would be adrenal steroid<br />
suppression and general excessive catabolism observed with high dose and
256 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
Figure 19.1 Classification <strong>of</strong> endocrine toxicity (Atterwill and Flack, 1993).<br />
prolonged use <strong>of</strong> glucocorticoids. In addition, the use <strong>of</strong> thyromimetic agents<br />
which may suppress pituitary thyrotroph function.<br />
Class 3 Effects which could not have been predicted from the<br />
pharmacology <strong>of</strong> the compound. This group can be subclassified into: (a)<br />
Effects which are direct or primary actions on an endocrine gland.<br />
Examples <strong>of</strong> this might be the action <strong>of</strong> ketoconazoles on adrenal and<br />
testicular function and the action <strong>of</strong> alloxan and streptozocin on the β-cell<br />
<strong>of</strong> the pancreas; and (b) effects on endocrine glands which are indirect or<br />
secondary to changes in other organs or control mechanisms<br />
(homeostasis). Examples here would be the actions <strong>of</strong> phenobarbitone and<br />
PCBs on the rat thyroid and the effects <strong>of</strong> lactose and polyols on the rat<br />
adrenal medulla.<br />
Class 4 Effects which cannot be predicted from pre-clinical studies<br />
because <strong>of</strong> idiosyncratic effects on the endocrine system.<br />
As indicated by Baylis and Tunbridge (1985) the adverse effects <strong>of</strong><br />
pharmaceutical agents on the endocrine system are generally due to normal
C.K.ATTERWILL AND S.P.AYLWARD 257<br />
or exaggerated pharmacological responses, that is Classes 1 and 2.<br />
Furthermore, it appears that endocrine toxicology can be detected reliably<br />
in pre-clinical studies. What is essential is that Class 3 toxic endocrine effects<br />
are classified appropriately. Class 3a effects would be expected to be seen<br />
across species, whereas there are numerous examples where Class 3b<br />
effects appear to be species-specific. There seems to be a paucity <strong>of</strong> clear-cut<br />
examples for Class 4 effects. There are several examples <strong>of</strong> compounds<br />
which have idiosyncratic immunotoxicological effects in susceptible humans<br />
which cannot be predicted from pre-clinical studies.<br />
Incidence <strong>of</strong> thyroid toxicity and tumours <strong>of</strong> the thyroid<br />
Most information on incidence is derived from pharmaceutical toxicity<br />
databases. Such data from Ribelin (1984) suggest that the endocrine system<br />
<strong>of</strong> the rats is particularly sensitive to toxicity from xenobiotics. This is also<br />
supported by Heywood (1984) in which he examined the target organ<br />
toxicity for 42 pharmaceutical compounds in the rat and dog. The<br />
endocrine system <strong>of</strong> the rat was only second to the liver as the most<br />
frequently affected target organ (38 per cent liver, 31 per cent endocrine).<br />
Ribelin (1984) reported that the most frequent endocrine lesion occurs in<br />
the adrenals followed by the testes but this analysis was conducted on<br />
chemicals and pharmaceuticals, and the data indicated that it was the<br />
cortical layers <strong>of</strong> the adrenal that were being predominantly effected,<br />
suggesting that the adrenal changes may be reflecting general stress<br />
responses rather than direct adrenal gland toxicity.<br />
In another analysis conducted in conjunction with the Centre <strong>of</strong><br />
Medicines Research this area was further explored. <strong>Toxicology</strong> data on<br />
124 compounds (all pharmaceuticals) were analysed. Just under 50 per<br />
cent (61/124) <strong>of</strong> these compounds have effects on one or more endocrine<br />
glands. Similar to Ribelin (1984) the adrenals were the most frequently<br />
affected, followed by the testes and the thyroid. An extensive survey <strong>of</strong> the<br />
different types <strong>of</strong> thyroid toxicity for both pharmaceutical agents and<br />
industrial chemicals was presented by Atterwill et al. (1993).<br />
Perturbation <strong>of</strong> thyroid function<br />
Thyroid function can be perturbed by agents affecting a number <strong>of</strong><br />
processes involved in the regulation <strong>of</strong> the hypothalamic-pituitary-thyroidliver<br />
(H-P-T-L) axis (Figure 19.2). These agents can affect function directly<br />
by interacting with thyroid cell receptors on their intracellular transduction<br />
mechanisms (see Figure 19.3). Alternatively thyroid function may be<br />
altered indirectly by agents affecting thyroid hormone metabolism and/or<br />
distribution—this event being followed by the release <strong>of</strong> thyrotrophic<br />
factors, or by xenobiotic-mediated alterations in the release <strong>of</strong> these factors
258 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
Figure 19.2 Hypothalamic-pituitary-thyroid-liver (HPTL) axis.<br />
themselves from the hypothalamus or pituitary gland (see also review by<br />
Cavalieri and Pitt-Rivers (1981) and Atterwill et al. (1993)). For example,<br />
thyromimetics such as L-T 3 or D-T 3 can cause pituitary atrophy and<br />
thyroid ‘shutdown’ and toxicity due to suppressive effects on the<br />
thyrotrophs in the adenohypophysis (Atterwill, 1988).
Figure 19.3 Thyroid follicular epithelial cell function.<br />
C.K.ATTERWILL AND S.P.AYLWARD 259<br />
Control <strong>of</strong> thyroid function and pathobiology <strong>of</strong> thyroid<br />
lesions<br />
Xenobiotic toxic effects on the hypothalamic-pituitarythyroid-liver<br />
(H-P-T-L) axis<br />
The control <strong>of</strong> mammalian thyroid follicular function is shown<br />
schematically in Figure 19.2 together with the points at which agents may<br />
perturb function and cause toxicity and thyroid lesions. The most common<br />
classes <strong>of</strong> agent affecting function are discussed at the different loci in<br />
terms <strong>of</strong> the categories (1–4) <strong>of</strong> endocrine toxicity (Atterwill and Flack,<br />
1993).<br />
The various control mechanisms and factors influencing hormone<br />
synthesis, distribution and metabolism can be summarised as follows:<br />
thyroid hormone (T 3 and T 4) synthesis and secretion from thyroid gland<br />
are controlled by thyroid stimulating hormone (TSH) released from the<br />
pituitary gland. This in turn is under control by hypothalamic thyrotrophin<br />
releasing hormone (TRH) and circulating levels <strong>of</strong> the thyroid hormones.<br />
Thyroid hormones exist in the circulation in the free (free T 4 (FT 4) and<br />
free T 3 (FT 3)) and protein-bound forms (approx 99 per cent <strong>of</strong> total T 4-<br />
TT 4 and total T 3-TT 3) and it is the FT 3 hormone produced by deiodination<br />
from T 4 which has both a physiological action on T 3 nuclear receptors in<br />
target tissues and influences pituitary TSH output.
260 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
Protein-binding <strong>of</strong> the hormone in the circulation takes place on several<br />
moieties, the pr<strong>of</strong>iles <strong>of</strong> which vary between different species (Dohler et al.,<br />
1973). The major proteins are thyroglobulin (TBG), thyroid binding<br />
pre albumin (TBPA) and albumin. Free and bound T 3 and T 4 are in<br />
dynamic equilibrium in the circulation and because <strong>of</strong> differences in affinity<br />
for the binding proteins there is more T 3 in the unbound form (approx 0.4<br />
per cent <strong>of</strong> total) than T 4 (approx 0.04 per cent <strong>of</strong> total).<br />
Thyroid hormone synthesis takes place at the apical membrane <strong>of</strong> the<br />
polarised follicular epithelial cells and depends on TSH-stimulated active<br />
iodide uptake under the influence <strong>of</strong> many extracellular and intracellular<br />
factors and signals (see Figure 19.3). Iodide is oxidised by peroxidase<br />
enzymes which mediate the incorporation <strong>of</strong> iodine into the tyrosyl<br />
residues <strong>of</strong> the colloidal glycoprotein, thyroglobulin. Colloid-bound<br />
thyroid hormone is stored in the follicular lumen and is released into the<br />
circulation by lysosomal action on the colloidal complex following<br />
endocytosis <strong>of</strong> colloid droplets into the cells. The incorporation and<br />
organification <strong>of</strong> iodide into thyroid hormone within thyroid follicles and<br />
the toxicological effects <strong>of</strong> xenobiotics can be studied in vivo using the<br />
‘perchlorate-discharge test’ (Atterwill et al., 1987) or in vitro using cultured<br />
thyrocytes (see Figure 19.4) from different species (Atterwill and Fowler,<br />
1990).<br />
Catabolic metabolism <strong>of</strong> the hormonal products <strong>of</strong> the thyroid gland, FT 3<br />
and FT 4, is achieved via two major pathways, deiodination and<br />
conjugation (glucuronidation and sulphation yielding a more water-soluble<br />
product for biliary excretion) and one minor route, deamination<br />
(decarboxylation, see Figure 19.9 later).<br />
The metabolic fate <strong>of</strong> thyroxine relies predominantly on its deiodination<br />
to T 3, only 20 per cent <strong>of</strong> circulating T 3 (the thyromimetic) is secreted by<br />
the thyroid (Engler and Burger, 1984). The remaining 80 per cent is derived<br />
from the deiodinative conversion <strong>of</strong> thyroxine (FT 4) to T 3. Deiodination<br />
can occur at several sites—the ones <strong>of</strong> major importance from the clearance<br />
aspect being liver and kidney whilst pituitary deiodination is essential for<br />
controlling responsiveness to circulating FT 4 levels. The deiodinases exist<br />
as three iso-zymes: Type I (5′-D; localised in liver, kidney, thyroid and<br />
central nervous system (CNS) tissue; and is propylthiouracil (PTU)<br />
sensitive); Type II (5'-D; localised exclusively in the CNS, brown adipose<br />
tissue, and pituitary; PTU-insensitive); and Type III (5-D, CNS; PTU<br />
insensitive) and have different affinities for T 4, different maturational<br />
patterns and different compensatory responses to hypothyroidism (see<br />
Kohrle et al., 1987).<br />
Bastomsky (1973), using Gunn rats, congenitally jaundiced due to UDPglucuronyl<br />
transferase deficiency (including T 4-glucuronyltransferase)<br />
demonstrated that the rate limiting step in hepatic thyroxine clearance (via
C.K.ATTERWILL AND S.P.AYLWARD 261<br />
the biliary excretion pathway) is the formation <strong>of</strong> the glucuronic acid<br />
conjugate, T 4-glucuronide.<br />
Hepatic conjugation, either sulphation (preferring FT 3) or<br />
glucuronidation (preferring FT 4) yields a more water soluble product,<br />
excreted in the bile/ biliary duct is a major pathway in T 4 excretion.<br />
Further deiodination via the deiodinase group <strong>of</strong> enzymes <strong>of</strong> T 4 T 3conjugates,<br />
T 4-amines and T 3 to thyromimetically inactive iodothyronines<br />
(rT 3 Triace, Tetrace, T 2 and T 1; see Figures 19.8 and 19.9) plays an<br />
important part in the T 4/T 3 biotransformation cascade and completes the<br />
thyroid hormone metabolic pr<strong>of</strong>ile.<br />
The key factor in maintaining correct thyroid follicular capability is an<br />
appropriate TSH output to TRH stimulation alongside circulating levels <strong>of</strong><br />
FT 4. Perturbation <strong>of</strong> this homeostatic control results in a classical thyroid<br />
response. The initial thyroid responses to increasing TSH levels are<br />
follicular cell hypertrophy, loss <strong>of</strong> colloid and vascular dilatation. In<br />
conventional animal toxicology studies performed for regulatory<br />
authorities one <strong>of</strong> the first indices <strong>of</strong> thyrotoxicity, therefore, is the<br />
observation <strong>of</strong> altered thyroid histopathology, primarily as follicular cell<br />
hypertrophy and/or diffuse hyperplasia, <strong>of</strong>ten leading to focal hyperplasia,<br />
thyroid adenomas and adenocarcinomas in longer term toxicity studies<br />
after longer term exposure.<br />
Pathobiology <strong>of</strong> thyroid follicular cell hyperplasia and<br />
neoplasia<br />
Thyroid neoplasia (see Figure 19.5) develops predictably in experimental<br />
species exposed to any procedure inducing prolonged and excessive TSH<br />
secretion (for example, the administration <strong>of</strong> chemical goitrogens, chronic<br />
iodine deficiency or subtotal thyroidectomy) although humans and mouse<br />
appear to be more resistant to TSH-induced thyroid neoplasia than rat.<br />
The number <strong>of</strong> cytogenetic abnormalities within the thyroid epithelium<br />
increases with duration <strong>of</strong> excess TSH exposure, with follicular cell<br />
hyperplasia potentially leading to neoplasia. The histopathological<br />
sequence <strong>of</strong> events is as follows (see Zbinden, 1987): following<br />
hypertrophy <strong>of</strong> the follicular epithelial cells focal hyperplasias appear in the<br />
gland which are distinct areas <strong>of</strong> papillary growth with enlarged epithelia.<br />
As these foci continue to grow they form nodules partly surrounded by<br />
collagenous fibres. These lesions are transition states between focal<br />
hyperplasias and adenomas. Adenomas are larger nodules that compress<br />
the surrounding tissue and have a distinct capsule. Follicular<br />
microcarcinomas (characterised by irregular gland-like structures,<br />
basophilia and nuclear crowding) may appear in some nodules. Larger<br />
carcinomas usually retain a follicular structure but sometimes consist <strong>of</strong><br />
solid sheets <strong>of</strong> polymorphous cells (Zbinden, 1987).
262 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
Figure 19.4 Cultured porcine thyrocytes in vitro. These scanning electron<br />
micrographs show (a) individual cultured ‘inverted’ follicles, and (b) individual<br />
follicular epithelial cells at higher magnification (apical membrane facing upwards)<br />
displaying TSH-stimulated microvilli.
C.K.ATTERWILL AND S.P.AYLWARD 263<br />
Figure 19.5 Factors influencing the development <strong>of</strong> thyroid neoplasia.<br />
Two consecutive processes are thought to occur in the development <strong>of</strong><br />
thyroid tumours; first, initiation which occurs quickly and is irreversible<br />
and secondly, promotion which occurs slowly and is reversible (and for<br />
which cell proliferation may be a necessary but not sufficient condition).<br />
Initiators may be ionising radiation, chemical/biological agents or genetic<br />
factors, with TSH acting as a promotion agent. Spontaneous thyroid<br />
follicular cell tumours arise from unknown aetiologies and factors and ageinduced<br />
changes in the cell membrane, growth factor and signaltransduction<br />
mechanisms may be involved. Small subpopulations <strong>of</strong> hyperreactive<br />
epithelial cells retaining the high replication rate <strong>of</strong> the foetal stage<br />
have been identified (Peter et al., 1982; Smeds et al., 1987) which may<br />
enter clonal expansion following only slight elevations in TSH (such as<br />
those in handled or stressed animals) or other growth factors, leading to<br />
spontaneous nodular goitres (see Zbinden, 1987).<br />
Many experimental studies have confirmed the key role <strong>of</strong> TSH as a<br />
stimulator <strong>of</strong> thyroid growth. In a rat thyroid cell line (FRTL-5) TSHstimulated<br />
growth <strong>of</strong> the cells was found to be associated with a marked<br />
increase in c-fos and c-myc oncogene expression (Colletta et al., 1986).<br />
Another example <strong>of</strong> the tumour promoting capacity <strong>of</strong> TSH is given by
264 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
studies where rats were given carcinogens such as N-methyl-N-nitrosourea<br />
(MNU) and then phenobarbital, or put on an iodine deficient diet. These<br />
treatments cause an early and increased incidence <strong>of</strong> thyroid follicular<br />
lesions and tumour formation (Hiasa et al., 1982; Oshima and Ward, 1984).<br />
The duration <strong>of</strong> exposure to high circulating TSH concentrations is also<br />
important in that intermittent administration <strong>of</strong> chemical goitrogens with<br />
TSH ‘normalisation’ does not appear to lead to follicular neoplasias.<br />
An elaborate series <strong>of</strong> studies have shown that a sustained elevation <strong>of</strong><br />
serum TSH in the rat leads to three phases <strong>of</strong> thyroid growth (Figure 19.6):<br />
(1) a phase <strong>of</strong> rapid growth lasting 1–2 months, followed by (2) a plateau<br />
phase <strong>of</strong> 3–6 months (growth desensitising mechanism (GDM) limiting<br />
epithelial cell mitotic response), followed eventually (3) by the appearance<br />
<strong>of</strong> multiple follicular cell tumours (loss <strong>of</strong> GDM; see Wynford-Thomas et<br />
al., 1982; Stringer et al., 1985; Smith et al., 1986). The reversibility <strong>of</strong> TSHinduced<br />
thyroid focal hyperplasia will evidently depend, therefore, on the<br />
stage during these ‘timed’ cellular changes in the first 6 months at which<br />
the TSH stimulus is withdrawn. Once the GDM is non-operative<br />
reversibility is not possible.<br />
Tumour progression seems to occur by a multi-stage process involving<br />
clonal ‘expansion’ and naturally occuring clones <strong>of</strong> cells have been<br />
demonstrated with high intrinsic proliferation potential in the mouse<br />
thyroid gland (Smeds et al., 1987), perhaps helping to explain the focal<br />
nature <strong>of</strong> hyperplastic and neoplastic lesions. The loss <strong>of</strong> a GDM within<br />
the follicular cells appears to be accompanied by an altered dependence or<br />
sensitivity to certain growth factors as well as the possible loss <strong>of</strong> an antioncogene<br />
which limits the follicular cells’s growth response to TSH. For<br />
example, the growth <strong>of</strong> normal cultured human thyroid cells requires TSH<br />
and insulin-like growth factor 1 (IGF1) in combination, whereas cells from<br />
adenomatous tissue in vitro proliferate in response to either TSH or IGF<br />
independently (Williams et al., 1987). This is due to the acquisition <strong>of</strong><br />
autocrine production <strong>of</strong> IGF 1 by the tumour cells themselves (see Thomas<br />
and Williams, 1991). Since the differentiation and growth <strong>of</strong> thyrocytes<br />
under TSH is regulated by cyclic-AMP-dependent mechanisms<br />
(Figure 19.2), tissue hyperplasia and hyperthyroidism might be expected to<br />
result when activation <strong>of</strong> the adenyl cyclase-cAMP cascade becomes<br />
unregulated. This can occur, for example, when somatic mutations impair<br />
the GTPase activity <strong>of</strong> G-protein coupled reactors, which may thus behave<br />
as proto-oncogenes. Such a mechanism is probably responsible for the<br />
development <strong>of</strong> a minority <strong>of</strong> monoclonal hyperfunctioning thyroid<br />
adenomas (Parma et al., 1993) (these also result in a silencing <strong>of</strong> normal<br />
thyroid function in extra-adenomatous tissue). Other non-genotoxic<br />
factors, such as agents affecting patterns <strong>of</strong> DNA methylation when<br />
coupled with a growth stimulus, should also be given consideration when
C.K.ATTERWILL AND S.P.AYLWARD 265<br />
Figure 19.6 Pathobiology <strong>of</strong> thyroid tumorigenesis.
266 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
attempting to define mechanisms in thyroid carcinogenesis (Thomas and<br />
Williams, 1992).<br />
In summary, there is, therefore, good evidence that sustained TSH drive<br />
to the thyroid gland can lead to a de-regulation <strong>of</strong> thyroid function. When<br />
investigating xenobiotic or drug-induced thyroid tumour formation, the<br />
mechanisms whereby TSH drive is increased can be understood by<br />
undertaking a series <strong>of</strong> experimental studies using in vitro and in vivo<br />
techniques. Having delineated the mechanism <strong>of</strong> the thyrotoxic effect it<br />
may then be possible to determine whether a particular drug or compound<br />
elicits a similar response in different species (including humans) and to<br />
investigate the dose-response relationship for this effect.<br />
Investigative toxicological studies and examples <strong>of</strong><br />
xenobiotics causing thyroid toxicity via the H-P-T-L axis<br />
Introduction<br />
Atterwill et al., (1993) give extensive examples <strong>of</strong> both pharmaceutical and<br />
industrial compounds causing thyroid toxicity via the five main sites along<br />
the H-P-T-L axis as shown in Figure 19.7 and readers should refer to this<br />
for further and more detailed information. In this chapter, three <strong>of</strong> these<br />
five thyroid toxicity loci are described in relation to the endocrine effects<br />
produced, industrial xenobiotic examples, and investigative in vivo and in<br />
vitro tests to delineate mechanisms and species-specific effects. This<br />
information is further summarised in Figure 19.8.<br />
In terms <strong>of</strong> industrial compounds the most frequently cited examples<br />
causing thyroid toxicity appear to be in the categories <strong>of</strong>: (i) those<br />
potentially affecting the plasma protein binding <strong>of</strong> thyroid hormones—for<br />
example, the nitrile herbicide, ioxynil (Ogilvie and Ramsden, 1988); (ii)<br />
those acting directly on the thyroidal peroxidase enzyme as goitrogens, and<br />
blocking thyroid hormone synthesis and secretion—for example, the coal<br />
derived hydroxyphenol products (Lindsay et al., 1992); and (iii) those<br />
affecting the hepatic metabolism and elimination T 3 and T 4—for example,<br />
compounds such as β-naphth<strong>of</strong>lavone, PCBs and alachlor (Ogilvie and<br />
Ramsden, 1988). Tables 19.1–19.3 show examples <strong>of</strong> these three class<br />
effects, compounds producing the effects and some <strong>of</strong> the range <strong>of</strong><br />
investigative tests currently available.<br />
in vivo and in vitro studies <strong>of</strong> xenobiotics acting on the<br />
hepatic metabolism and clearance <strong>of</strong> thyroxine<br />
There is a growing list <strong>of</strong> agents, both pharmaceutical and industrial<br />
xenobiotics, which act in rodents by interfering with thyroid hormone
Figure 19.7 Toxicological loci in H-P-T-L axis.<br />
C.K.ATTERWILL AND S.P.AYLWARD 267<br />
metabolism, hepatic elimination and thus circulating TSH levels (see also<br />
Capen and Martin, 1989; McClain, 1989; Atterwill et al., 1993). The<br />
xenobiotics include phenobarbital (McClain, 1989), β-naphth<strong>of</strong>lavone<br />
(Johnson et al., 1993), the polychlorinated biphenyls (Bastomsky, 1974),<br />
diproteverine (a calcium antagonist; Flack et al., 1989), SC37211 (a Searle<br />
imidazole antimicrobial (Comer et al., 1985), L649923 (a leukotriene D 4<br />
antagonist; Saunders et al., 1988), a novel oxyacetamide-FOE 5043<br />
(Christenson et al., 1993), alachlor (Brewster et al., 1993), PCNB<br />
(pentachloronitrobenzene; Story et al., 1993), and hexachlorobenzene<br />
(Ogilvie and Ramsden, 1988).<br />
Most <strong>of</strong> these compounds have thus far been assumed to act in vivo via<br />
the induction <strong>of</strong> hepatic uridine diphosphate glucuronosyltransferase (UDP-<br />
GT) in the rat, with species-specific formation <strong>of</strong> thyroid tumours in<br />
carcinogenicity studies (see McClain, 1989). Indeed many <strong>of</strong> the<br />
compounds, including phenobarbital do lead to increased hepatic UDP-GT<br />
activity and appearance <strong>of</strong> glucuronidated T 4 in the bile, sometimes with<br />
elevated bile flow rates (see McClain, 1989). However, others such as the
268 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
Figure 19.8 Investigative tests on H-P-T-L Axis.<br />
pharmaceutical temelastine increase predominantly the clearance <strong>of</strong> free<br />
T 4, though the bile product is not in conjugate form (Poole et al., 1989,<br />
1990). Other compounds such as the food dye FD&C Red No 3 (Capen<br />
and Martin, 1989) are able to lower circulating triiodothyronine (T 3) by<br />
altered deiodination suggesting the further existence <strong>of</strong> alternative<br />
mechanisms. Furthermore, not all chemicals inducing hepatic neoplasia in<br />
rodents cause thyroid neoplasia (McClain, 1989).<br />
We and others have reported that two SK&F histamine antagonists,<br />
temelastine (SK&F 93944) and lupitidine (SK&F 93479) produce ratspecific<br />
thyroid lesions via perturbation <strong>of</strong> the hepatic locus (Atterwill et
Table 19.1<br />
Table 19.2<br />
C.K.ATTERWILL AND S.P.AYLWARD 269<br />
al., 1989). Increased thyroxine clearance from the circulation, followed by<br />
elevated TSH ‘drive’ and increased thyroid follicular cell growth were<br />
observed. These com pounds act rapidly within minutes—hours <strong>of</strong> in vivo<br />
drug administration and are apparently able to increase the accumulation <strong>of</strong><br />
T 4 directly in vitro by cultured rat hepatocytes (Atterwill et al., 1989; Poole<br />
et al., 1990). Phenobarbital appears to share this property in vitro
270 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
Table 19.3<br />
(Aylward et al., 1994) and increases the accumulation <strong>of</strong> thyroxine in<br />
treated rat liver in vivo (Oppenheimer et al., 1968).<br />
Thyroxine transport (Figure 19.9) is regulated by specific components<br />
located within the plasma membrane in various cell types including<br />
fibroblasts and hepatocytes and is an important prerequisite for both<br />
hormone metabo lism and nuclear hormonally-mediated events (Pliam and<br />
Goldfine, 1977; Krenning et al., 1981; Blondeau, 1986). Investigations<br />
indicate that there exist two distinct transport systems specific to<br />
thyroxine: a high-affinity, low capacity, energy-dependent ATP-ase linked<br />
transport system and a low-affinity, high capacity transport mechanism<br />
(Sorimachi and Robbins, 1978; Krenning et al., 1981, 1983; Blondeau,<br />
1986; Rao, 1991).<br />
Many <strong>of</strong> the compounds listed as indirect carcinogens in rat (due to an<br />
ability to induce liver microsomal enzymes and increase glucuronidated<br />
thyroxine elimination in the bile) have been shown to increase hepatic UDP-<br />
GT activity or cause liver hypertrophy indicative <strong>of</strong> following repeated<br />
dosing (Comer et al., 1985; McClain, 1989; Johnson et al., 1993).<br />
However, there have been no attempts to provide a definitive link between<br />
UDP-GT induction and thyroid pathology or to prove a primary<br />
endocrinological effect via UDP-GT. Our previous work with temelastine<br />
in the rat in vivo (Atterwill et al., 1989) was able to demonstrate that the<br />
increased clearance <strong>of</strong> T 4 from the rat circulation appeared within a few<br />
hours <strong>of</strong> a single compound dosing (Atterwill et al., 1989). Even
Figure 19.9 Hepatic events leading to hormone elimination.<br />
C.K.ATTERWILL AND S.P.AYLWARD 271<br />
phenobarbital is able to increase the thyroxine clearance in a relatively<br />
short timespan (Atterwill, unpublished observations). New in vitro data are<br />
not inconsistent with the time course <strong>of</strong> the in vivo phenomena where<br />
enhanced T 4 hepatocytic accumulation by cultured rat hepatocytes<br />
following compound exposure occurred as early as 60–90 min after<br />
exposure (Aylward et al., 1994). There was no membrane cytotoxic effect<br />
<strong>of</strong> the compounds at the threshold concentrations producing these effects in<br />
vitro. This shows a potential rapid direct effect <strong>of</strong> the xenobiotics on
272 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
Table 19.4 Species and energy dependence <strong>of</strong> enhanced thyroxine accumulation in<br />
vitro<br />
Key: ↑T 4, increase; ↓T 4, decrease; ↔T 4, no change; Temperature? ATP?,<br />
temperature/ATP dependent; NA, not applicable.<br />
hepatocellular T 4 accumulation. The correlation between in vivo—in vitro<br />
species-specific toxicological effects is also evident (Figure 19.10 and<br />
Table 19.4). One <strong>of</strong> the features <strong>of</strong> temelastineinduced thyroid toxicology<br />
in vivo was the apparent species-specificity to the rat (Atterwill et al.,<br />
1989; Poole et al., 1989). Temelastine-mediated thyroid follicular<br />
hypertrophy and hyperplasia was not observed in dog, mouse or monkey<br />
following temelastine treatment (Figure 19.10). In vitro, no enhanced<br />
thyroxine accumulation in response to temelastine or phenobarbital was<br />
observed in guinea pig or dog hepatocytes (Aylward et al., 1994). In<br />
support <strong>of</strong> these findings, it has been demonstrated that the guinea pig is<br />
insensitive to thyroid pathological changes after phenobarbital or βnaphth<strong>of</strong>lavone<br />
administration in vivo (Johnson et al., 1993; Wyatt et al.,<br />
1993). In support <strong>of</strong>, and as an extension <strong>of</strong> these findings, we now present<br />
important new findings to demonstrate conclusively that some <strong>of</strong> the rapid<br />
‘effectors’ <strong>of</strong> thyroid toxicity via the liver, such as temelastine, do so<br />
independently <strong>of</strong> a primary action on UDP-GT, whereas other cytochrome<br />
P450 inducers such as phenobarbital may have a combined effect. This<br />
work was carried out using hepatocytes prepared from UDP-GT system<br />
deficient Gunn rats.<br />
Studies on Gunn rat hepatocytes in vitro<br />
Hepatocytes were prepared from the normal or Gunn rat (deficient in UDP-<br />
GT isozymes conjugating thyroxine) and exposed to either temelastine or<br />
phenobarbital (2 or 20 µM) for 3 h as before (Aylward et al., 1994). The<br />
results show (Figure 19.11) that whereas temelastine was able to enhance<br />
thyroxine accumulation in both types <strong>of</strong> hepatocytes, phenobarbital only<br />
produced alterations in hormone accumulation in normal cells, supporting
C.K.ATTERWILL AND S.P.AYLWARD 273<br />
Figure 19.10 Effect <strong>of</strong> temelastine on 125I-T4 clearance (from Atterwill et al.,<br />
(1989)).
274 ENDOCRINE TOXICOLOGY OF THE THYROID<br />
Figure 19.11 Effect <strong>of</strong> temelastine and phenobarbital on thyroxine accumulation in<br />
vitro by hepatocytes from control and Gunn rats.<br />
earlier in vivo findings where phenobarbital was toxicologically inactive in<br />
this strain <strong>of</strong> rat (Bastomsky, 1973).
C.K.ATTERWILL AND S.P.AYLWARD 275<br />
Conclusions<br />
The observations now lend further and strong support to the hypothesis<br />
that indirect xenobiotic-induced thyroid toxicology can arise from direct<br />
effects on hepatic membrane-located thyroxine transport proteins. It also<br />
suggests that the species-specificity <strong>of</strong> this toxic effect in vivo <strong>of</strong> some<br />
xenobiotics may be attributed to actual species differences in the sensitivity<br />
<strong>of</strong> these hepatic carriers to the compounds and not simply or primarily to<br />
changes in T 4 glucuronidation via UDP-GT induction. For the first time we<br />
have demonstrated the usefulness <strong>of</strong> Gunn rat hepatocytes in vitro for<br />
discriminating between the two ‘hepatic subclasses’ <strong>of</strong> xenobiotics causing<br />
thyroid toxicity in rodents.<br />
A number <strong>of</strong> practical in vivo and in vitro investigative tests are now<br />
available for delineating mechanisms <strong>of</strong> thyroid toxicity along the H-P-T-L<br />
axis, and which also provide screening tools for examining chemical series<br />
<strong>of</strong> potentially toxic molecules: (i) Direct block <strong>of</strong> thyroid function via<br />
peroxidase inhibition can be measured in vivo by the perchlorate discharge<br />
test (Atterwill et al., 1987); (ii) it can also be measured in vitro using<br />
cultured thyrocytes (Atterwill and Fowler, 1990); (iii) indirect effects on<br />
hepatic thyroxine clearance can be assessed in vivo (Atterwill et al., 1989);<br />
or (iv) in vitro using cultured hepatocytes from different species or Gunn<br />
rat (Aylward et al., 1994). Effects on receptors at the hypothalamic and<br />
pituitary levels can also now be studied extensively using both in vivo and<br />
in vitro approaches (Buckingham and Gillies, 1993). This battery <strong>of</strong><br />
technology now available will greatly advance the mechanistic<br />
understanding and screening <strong>of</strong> thyroid endocrine toxicants.<br />
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20<br />
Testing and Evaluation for Reproductive Toxicity<br />
ANTHONY K.PALMER<br />
Huntingdon Research Centre, Huntingdon<br />
Introduction<br />
Most <strong>of</strong> the presentations at this meeting refer to high level, scientific<br />
investigations <strong>of</strong> one or two, highly important, high production volume<br />
chemicals, for which an adverse effect has been demonstrated. They are<br />
studies <strong>of</strong> characterisation, because they elaborate on known effects using a<br />
wealth <strong>of</strong> available information. But, how were the adverse effects <strong>of</strong> these<br />
few substances first discovered, what were the initial clues? Sadly, for<br />
many, the observation <strong>of</strong> adverse effects in humans was the trigger to<br />
intensive investigations, which is akin to ‘shutting the stable door after the<br />
horse has bolted’.<br />
This presentation is concerned with detecting effects <strong>of</strong> substances for<br />
which little or no information is available and, preferably, before they<br />
cause harm to humans. This requires a different kind <strong>of</strong> science, for which<br />
the main asset is the ability to predict, with reasonable accuracy, possible<br />
activity from minimal information. It requires wide experience and a<br />
balance between imagination and pragmatism. These attributes are<br />
especially important for toxicity to reproduction, which triggers instinctive<br />
reactions in even the coolest and most objective scientist.<br />
Identifying the cause <strong>of</strong> adverse effects on human reproduction has long<br />
been surrounded by controversy and uncertainty. In respect <strong>of</strong> the<br />
evaluation <strong>of</strong> substances for reproductive toxicity this state <strong>of</strong> affairs seems<br />
likely to persist for years to come. The main obstacle to any attempt to<br />
rationalise the situation is that any discussion on evaluation almost<br />
inevitably gravitates to the black hole <strong>of</strong> regulatory guidelines. All<br />
guidelines are flawed because science fact is compromised by bureaucracy<br />
and science fiction. For many reasons, but especially the unwillingness <strong>of</strong><br />
any establishment to change the status quo, guidelines provide the worst<br />
starting point for developing a strategy for evaluation.<br />
Most guidelines are concerned only with methods for gathering specified<br />
information. On its own this information (on hazard) is insufficient and<br />
needs to be supplemented with other information, from other sources, to
Table 20.1 Production volume triggers for industrial chemicals<br />
Table 20.2 EC Annex VII and VIII toxicity tests for industrial chemicals<br />
A.K.PALMER 281<br />
Notes:<br />
Tests for ecotoxicity are not included.<br />
Progression, by rote, from base set to level 2 would involve duplication <strong>of</strong> effort in<br />
several areas.<br />
The state <strong>of</strong> confusion regarding testing for toxicity to reproduction is portrayed by<br />
the failure to decide on requirements at base set level and the curious mixture <strong>of</strong> old<br />
and new terminology regarding tests.<br />
predict whether humans might be affected. Most guidelines are a watershed<br />
in a broader spectrum <strong>of</strong> testing and assessment. They represent a point at<br />
which it may be decided that the only way to gather more information is to<br />
take the final step <strong>of</strong> exposing humans. Exposure <strong>of</strong> humans provides the
282 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY<br />
Figure 20.1 Overlap <strong>of</strong> toxicity.<br />
only certain way to determine whether reproduction would be affected, but<br />
we need to do the best we can before taking the chance.<br />
The paradox in this is that exposure <strong>of</strong> humans is facilitated by failure to<br />
demonstrate toxicity in animals. But, lack <strong>of</strong> activity, being negative, cannot<br />
be proven, only presumed. To make this presumption investigations must<br />
be extensive and comprehensive to convey reasonable assurance that failure<br />
was not due to deficiencies in methodology.<br />
There is an exception. For industrial chemicals, testing <strong>of</strong> all substances<br />
to these criteria would be a monumental task, therefore, less stringent<br />
testing is allowed according to production volume, which serves as an<br />
approximation for the extent <strong>of</strong> the population likely to be exposed<br />
(Table 20.1). It is a strategy based on population risk, the downside <strong>of</strong><br />
which is an increased risk to the individual. The strategy takes advantage<br />
<strong>of</strong> the fact that, in a small population, the chance <strong>of</strong> identifying cause and<br />
effect is poor. At the base set level and level 1, as outlined by the EC, all<br />
tests are equivalent only to the voluntary preliminary studies conducted for
medicines, agrochemicals and food additives. They may be sufficient to<br />
detect potent toxicity but not comprehensive enough to allow presumption<br />
<strong>of</strong> the absence <strong>of</strong> hazard, especially as the base set does not include tests<br />
for reproductive toxicity (Table 20.2).<br />
As production volume increases, more extensive testing should be<br />
undertaken. Often, this has been neglected, prompting the development <strong>of</strong><br />
the OECD guidelines 421 and 422. These tests were intended to recover a<br />
situation that never should have arisen.<br />
A better approach?<br />
There is no question that evaluation for reproductive toxicity could be<br />
improved considerably. The question is whether industry and agencies are<br />
willing to do so. It would require a change <strong>of</strong> attitude in industry and<br />
agencies alike. Industry’s ‘passive avoidance’ <strong>of</strong> testing would need to be<br />
replaced by ‘active participation’. For a new substance the first step should<br />
be an integrated assessment <strong>of</strong> commercial prospects and potential toxicity<br />
over a broad spectrum (Figure 20.1). Early identification <strong>of</strong> ‘serious bad<br />
actors’, which tend to effect many systems, can save time and effort. Given<br />
the prognosis <strong>of</strong> problems ahead, it may be better to devote resources to<br />
finding safer alternatives, or to risk management, rather than to endless<br />
testing. For materials with a high commercial potential, the aim should be<br />
to get to full scale tests by the quickest route. Following the EC levels by rote<br />
is very inefficient since there is duplication with successive steps.<br />
Methods<br />
With an active participation policy, a much broader scope <strong>of</strong> methodology<br />
can and should be considered, ranging from searches for structure-activity<br />
relationships, through various in vitro methods, whole animal tests, wild<br />
life surveys and human surveys (Table 20.3). Due to time constraints I will<br />
concentrate on whole animal test systems.<br />
Structure-activity databases<br />
Structure and activity relationships are an obvious place to start any<br />
evaluation, despite the fact that currently available databases are far from<br />
perfect. Their reliability could be improved dramatically by adding unused<br />
information currently hidden in industry and agency archives.<br />
Entire mammalian tests<br />
A.K.PALMER 283<br />
Tests in entire mammals provide the only way <strong>of</strong> assessing what effects a<br />
substance may evoke in the complex, integrated and dynamic process <strong>of</strong>
284 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY<br />
Table 20.3 Methods for detecting effects on human reproduction<br />
reproduction. To detect the wide range <strong>of</strong> possible effects it is necessary to<br />
expose mammals to a substance from conception through sexual maturity.<br />
It is necessary to look for consequences <strong>of</strong> this exposure through at least<br />
one life cycle (Figure 20.2). This long observation period is required for<br />
detection <strong>of</strong> latent manifestations <strong>of</strong> developmental toxicity, such as those<br />
induced by lead, alcohol, diethylstilboestrol and other hormonally active<br />
substances. The only means <strong>of</strong> covering all these aspects is a two generation<br />
study or the equivalent in a combination <strong>of</strong> tests.<br />
Restricted test systems<br />
The use <strong>of</strong> lesser tests for industrial chemicals is a concession. Examination<br />
for some effects is omitted because they are not perceived to be important<br />
or because they would be difficult to detect, or because they occur very<br />
rarely. For example, first detection <strong>of</strong> effects in <strong>of</strong>fspring <strong>of</strong> second<br />
generations is rare so such activity does not have a high priority.<br />
With these restricted tests, emphasis should be on detecting effects and<br />
not on manipulating a no effect level. Tests that could be considered would<br />
include OECD 421 and 422, the OECD single generation study and the old<br />
FDA Segment I study for medicines. The latter two are restricted tests<br />
because they do not allow detection <strong>of</strong> latent manifestations <strong>of</strong><br />
developmental toxicity.<br />
The best return for effort is afforded by OECD 422, which combines<br />
examination for general or systemic toxicity, as well as reproductive<br />
toxicity. However, realising its potential requires an experienced laboratory<br />
team, the courage to modify the test and the conceptual ability to know<br />
how to interpret the results.
Figure 20.2 Cycle <strong>of</strong> life/reproduction.<br />
A.K.PALMER 285<br />
OECD 421 involves treatment <strong>of</strong> both sexes from about 2 weeks prior to<br />
mating through to termination, a few days after birth <strong>of</strong> <strong>of</strong>fspring<br />
(Figure 20.3). Assessment <strong>of</strong> male fertility is achieved in two parts. Males<br />
are paired with females for detection <strong>of</strong> effects unrelated to<br />
spermatogenesis, for example, effects on sexual behaviour, libido or<br />
ejaculation and functional maturation <strong>of</strong> sperm.<br />
For detecting effects on spermatogenesis, direct methods, particularly<br />
histopathological examinations <strong>of</strong> testes and epididymides are used. Sperm<br />
analysis (seminology) could be added, although it does not seem to be<br />
better than histopathology. These methods could be incorporated into<br />
systemic toxicity studies rather than in the reproduction study per se.<br />
In respect <strong>of</strong> fecundity, treatment and observations <strong>of</strong> females include<br />
most <strong>of</strong> those applied in full scale tests. An exception is the lack <strong>of</strong><br />
observations for delayed, post-natal manifestations. Detailed examination<br />
<strong>of</strong> foetuses for skeletal and s<strong>of</strong>t tissue abnormalities is not included. The<br />
potential for prenatal effects is deduced by observation <strong>of</strong> post-natal<br />
differences in numbers pregnant, litter size, litter and mean pup weight at<br />
birth and to day 4 post partum.<br />
As with any guideline, OECD 421 should be used with commonsense<br />
and flexibility. If pretesting prognosis suggests that prenatal effects are<br />
unlikely, extension <strong>of</strong> the study to weaning <strong>of</strong> the <strong>of</strong>fspring (Figure 20.3)<br />
provides added safeguards at little extra cost. Increase group size and it
286 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY<br />
Figure 20.3 OECD 421 priority selection test.<br />
Figure 20.4 Fertility and embryotoxicity.<br />
provides the equivalent <strong>of</strong> the OECD single generation study. Conversely,<br />
if pretesting prognosis suggests a high probability <strong>of</strong> prenatal effects,<br />
including induction <strong>of</strong> malformation, then females could be killed just<br />
before delivery and foetuses examined for structural defects (Figure 20.4).<br />
This provides the equivalent <strong>of</strong> a fertility and embryotoxicity study we will<br />
see again later.<br />
OECD guideline 422 simply adds to OECD 421, elements for assessment<br />
<strong>of</strong> systemic and neurotoxicity. For those who have never conducted such a<br />
test it seems impossibly complex, but it is neither as difficult to perform,<br />
nor to interpret, as is feared. Its rejection by EC Officialdom makes it an<br />
even better proposition, since there need be no inhibitions about modifying<br />
the design according to circumstances.<br />
Some brief examples <strong>of</strong> results that may be encountered with positive<br />
materials are illustrated by the examples <strong>of</strong> Carbendazim (metabolite <strong>of</strong><br />
Benomyl), DEHP, Cyclophosphamide and ethylene glycol methyl ether<br />
(EGME, 2-methoxyethanol). With Carbendazim (Table 20.4) macroscopic<br />
and microscopic examinations show unequivocal effects on testes and<br />
epididymides indicating an effect on spermatogenesis. An effect on females<br />
and <strong>of</strong>fspring is indicated by an increased duration <strong>of</strong> pregnancy, reduction<br />
in the number <strong>of</strong> females with live young and lower values for litter size,
Table 20.4 OECD 422: carbendazim, tabular summary<br />
Notes:<br />
a Malformations included hydrocephaly and misaligned tails.<br />
pp=post partum<br />
Bold type indicates treatment effects including macroscopic and microscopic<br />
changes in testes and epididymides (an effect on spermatogenesis), an increased<br />
duration <strong>of</strong> pregnancy, reduction in the number <strong>of</strong> females with live young and lower<br />
values for litter size, litter weight and mean pup weight. The dosage related pattern<br />
<strong>of</strong> response provides added emphasis, as does the observation <strong>of</strong> malformed<br />
foetuses.<br />
A.K.PALMER 287<br />
litter weight and mean pup weight. The dosage related pattern <strong>of</strong> response<br />
provides added emphasis, as does the observation <strong>of</strong> malformed foetuses.<br />
With DEHP (Table 20.5) an effect on spermatogenesis is evident at 2000<br />
mg kg −1 . Treatment at this dosage had to be withdrawn shortly after<br />
mating to avoid further mortalities <strong>of</strong> the more susceptible females. There<br />
is a marked reduction in the number <strong>of</strong> pregnancies. At lower dosages an<br />
increased duration <strong>of</strong> pregnancy, reduced litter size and litter weight,<br />
indicates an effect on the female and/or <strong>of</strong>fspring. The higher mean pup<br />
weights are consequent to the longer duration <strong>of</strong> pregnancy.<br />
With cyclophosphamide (Table 20.6) female deaths at 3 and 4.5 mg kg −1<br />
are attributable to treatment as, at 6.7 mg kg −1 , all females died. There was<br />
a reduction in the number <strong>of</strong> females with young, litter size, litter weight<br />
and mean pup weight, providing clear evidence <strong>of</strong> effects on the female and
288 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY<br />
Table 20.5 OECD 422: DEHP, tabular summary<br />
Notes:<br />
[] Treatment at 2000 mg kg −1 was withdrawn after mating (4 weeks <strong>of</strong> treatment)<br />
due to loss <strong>of</strong> condition and mortality <strong>of</strong> females.<br />
pp=post partum.<br />
Bold type indicates treatment effects on spermatogenesis at 2000 mg kg −1 . There is<br />
a marked reduction in the number <strong>of</strong> pregnancies. At lower dosages an increased<br />
duration <strong>of</strong> pregnancy, reduced litter size and litter weight, indicates an effect on<br />
the female and or <strong>of</strong>fspring. The higher mean pup weights are consequent to the<br />
longer duration <strong>of</strong> pregnancy.<br />
<strong>of</strong>fspring. No effects on spermatogenesis were reported but, if a dosage<br />
inducing effects on the male had been selected, all the females would have<br />
died.<br />
With EGME (Table 20.7), dosages were based on acute toxicity and<br />
limited repeat dose toxicity studies only. This provided a more realistic<br />
representation <strong>of</strong> the testing <strong>of</strong> a new substance. The first consequence was<br />
the occurrence <strong>of</strong> systemic toxicity at 500 and 1000 mg kg −1 . Treatment at<br />
the high dosages had to be withdrawn prior to mating, investigating<br />
recovery became a new objective. At 100 mg kg −1 pr<strong>of</strong>ound effects on<br />
spermatogenesis were evident. Some pregnancies were obtained, but no live<br />
young were born. For the high dosage groups, effects on males remained<br />
evident several weeks after withdrawal <strong>of</strong> treatment. The duration <strong>of</strong><br />
pregnancy was increased. A consequence <strong>of</strong> this is the higher mean pup<br />
weight. Values for numbers <strong>of</strong> implantations, live young and litter weight<br />
were lower than control values. So, not only did the test show the<br />
anticipated effects on reproduction, it also demonstrated that recovery <strong>of</strong>
Table 20.6 OECD 422: cyclophosphamide, tabular summary<br />
Notes:<br />
pp=post partum.<br />
Bold type indicates treatment effect. Female deaths at 3 and 4.5 mg kg −1 are<br />
attributable to treatment as, at 6.7 mg kg −1 , all females died. There was a reduction<br />
in the number <strong>of</strong> females with young, litter size, litter weight and mean pup weight.<br />
No effects on spermatogenesis were reported but, if a dosage inducing effect on the<br />
male had been selected, all the females would have died.<br />
females was slow. As far as I know this has not been mentioned in the<br />
extensive literature on EGME.<br />
These examples and others show that the OECD tests are capable <strong>of</strong><br />
detecting substances with marked effects on reproduction. With such<br />
results it would be foolhardy to consider higher level guideline tests. If the<br />
substance is not abandoned any further testing would require case by case<br />
designs to characterise the detected effects more completely.<br />
Full scale testing<br />
If pretesting prognosis suggests that a substance is unlikely to present<br />
problems <strong>of</strong> toxicity, fast track progression to level 2 testing should be<br />
considered (Table 20.2) to avoid unnecessary duplication <strong>of</strong> step by step<br />
testing. Expected tests include a two generation study in rats and an<br />
embryotoxicity study in rats and rabbits. In the harmonised guideline for<br />
medicines this same strategy is just one <strong>of</strong> several options and the same or<br />
greater flexibility should be made available for testing industrial<br />
chemicals.<br />
Tests for embryotoxicity<br />
A.K.PALMER 289<br />
An anomaly in this strategy is the specification for embryotoxicity studies<br />
in two species, when only one species is required for all other aspects <strong>of</strong>
290 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY<br />
Table 20.7 OECD 422: EGME, tabular summary<br />
Notes:<br />
[] Treatment at 500 and 1000 mg kg −1 withdrawn prior to mating due to loss <strong>of</strong><br />
condition and mortality. Animals in withdrawal phase.<br />
NE not examined<br />
pp=post partum<br />
Bold type indicates treatment effect. At 100 mg kg −1 pr<strong>of</strong>ound effects on<br />
spermatogenesis were evident. Some pregnancies were obtained, but no live young<br />
were born indicating effects on females and the conceptus.<br />
For the high dosage groups, effects on testes and epididymides remained evident<br />
several weeks after withdrawal <strong>of</strong> treatment. The duration <strong>of</strong> pregnancy was<br />
increased with a consequent increase in mean pup weight. Values for numbers <strong>of</strong><br />
implantations, live young and litter weight were reduced indicating slow recovery.<br />
This does not appear to have been mentioned in the extensive literature on EGME.<br />
reproductive toxicity. A more sensible strategy would be to identify a<br />
relevant species before testing. It is pointless to conduct a test in an<br />
unsuitable species and doubly pointless to conduct tests in two irrelevant<br />
species.<br />
The requirement for detailed examination <strong>of</strong> foetuses for abnormalities<br />
is based on an exaggerated perception <strong>of</strong> risk prompted by fear. For many<br />
reasons the risk <strong>of</strong> inducing abnormalities is extremely low. Those same<br />
reasons make detection by direct observation <strong>of</strong> malformations unreliable.
Table 20.8 Post-natal detection <strong>of</strong> prenatal effects<br />
A.K.PALMER 291<br />
Notes:<br />
Prenatal effects include any s<strong>of</strong>t tissue or skeletal changes (variants, anomalies or<br />
malformations) or altered foetal weight. Post-natal effects include reduced litter size<br />
at birth, reduced mean foetal weight or increased post natal mortality.<br />
In the few studies in which post-natal effects were not detected, prenatal changes<br />
were <strong>of</strong> a ‘minor’ nature (e.g. reduced ossification) and sometimes not conclusive.<br />
The lower number <strong>of</strong> litters reared in the Japanese Experiment 2 design would<br />
contribute to the slightly higher rate <strong>of</strong> ‘failures’.<br />
Where more serious prenatal effects such as the observation <strong>of</strong> malformations or<br />
prenatal death were observed a post-natal effect was always observed.<br />
The dimensions <strong>of</strong> the tests we conduct are too small and dosage regimes<br />
are contradictory to the basic principle that malformations are induced by<br />
application <strong>of</strong> a precise dosage at a precise time. Whilst direct observation<br />
<strong>of</strong> malformations is unreliable the saving grace is that induced<br />
malformations always occur within a wider spectrum <strong>of</strong> embryotoxicity.<br />
This wider spectrum provides a more reliable, if indirect, means <strong>of</strong><br />
detecting substances that might cause malformations. Also, these effects are<br />
important in their own right.<br />
Major effects such as altered <strong>of</strong>fspring weight and prenatal death can be<br />
observed postnatally. For example both EC Segment I and Japanese<br />
Experiment 2 studies include examinations for foetal abnormalities as well<br />
as postnatal observations. A survey <strong>of</strong> such studies shows that, when<br />
malformations were observed, a post-natal effect was always observed<br />
(Table 20.8). Surveys such as this should have been conducted or<br />
sponsored by agencies and industry before formulating guidelines. Why<br />
have they not done so?<br />
The obsession with abnormalities is based on the fear <strong>of</strong> another<br />
thalidomide. The great contradiction is that a considerable number <strong>of</strong> rat<br />
embryotoxicity studies with thalidomide failed to provided convincing<br />
demonstration <strong>of</strong> teratogenicity. Conversely, reproduction studies in rats<br />
provided unequivocal, post-natal evidence <strong>of</strong> effect in the form <strong>of</strong> a marked<br />
reduction in the number <strong>of</strong> females with young and a marked reduction in
292 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY<br />
Table 20.9 Thalidomide—rat reproduction studies<br />
Notes:<br />
Thalidomide was administered in the diet to provide a daily intake <strong>of</strong> 200 mg kg −1<br />
bodyweight from 60 days prior to mating (US FDA two litter test). Later studies<br />
demonstrated that the reduced percentage <strong>of</strong> females with young and the lower live<br />
litter size in females with young was associated with embryolethality.<br />
litter size <strong>of</strong> the few that had young. Two studies, each containing several<br />
matings showed the reproducibility <strong>of</strong> the results (Table 20.9).<br />
Two generation studies<br />
The emphasis on structural abnormalities detracts from examination for<br />
other, important and more likely manifestations <strong>of</strong> reproductive toxicity<br />
(Figure 20.5). For example, current and proposed test guidelines for nonmedicines<br />
lack procedures for detection <strong>of</strong> developmental neurotoxicity (or<br />
behaviour). This is a curious contradiction given the current fashion for<br />
investigation <strong>of</strong> adult neurotoxicity.<br />
For detecting other effects on reproduction, all regulatory versions <strong>of</strong> the<br />
two generation study leave something to be desired. Even more<br />
disappointing is that newer versions proposed by the US FDA, the US EPA,<br />
the EC and OECD have recycled many <strong>of</strong> the old flaws. Truly, there has<br />
been a great deal <strong>of</strong> activity but very little progress.<br />
All current and proposed guidelines continue to require a prolonged<br />
premating treatment period for the F0 or parent generation. The claim that<br />
it is necessary to treat males for a full spermatogenic cycle is pure science<br />
fiction. Spermatogenesis is not a cycle but a sequence <strong>of</strong> overlapping batch<br />
processes. At any one time, all stages <strong>of</strong> spermatogenesis are present and a<br />
short dosing period is sufficient to cause effects.<br />
To detect these effects, direct histopathological methods can be applied<br />
shortly after treatment. Direct methods are quicker and more certain than<br />
mating to females. Mating trials are inefficient and lack sensitivity due to
Figure 20.5 Manifestations <strong>of</strong> developmental toxicity<br />
A.K.PALMER 293<br />
the high sperm production capacity <strong>of</strong> animals compared with humans.<br />
Interim results from an ongoing survey show that, where data are<br />
available, direct methods are effective and that a combination <strong>of</strong> direct<br />
methods with a premating dosing time <strong>of</strong> 2 weeks or less is even more<br />
effective (Table 20.10). The combination compares favourably with<br />
prolonged premating treatment and mating. Why have agencies and<br />
industry failed to conduct or sponsor such surveys?<br />
For non-medicines, use <strong>of</strong> a prolonged premating treatment period for the<br />
parent generation is an unnecessary duplication as treatment is continued<br />
into the F1 generation; this cannot be mated until animals have reached<br />
sexual maturity. Using science facts, a more efficient design for a two<br />
generation study (Figure 20.6) would include the following features:<br />
– One control and 2–4 test groups with dosages set at 2–5 fold descending<br />
intervals from a high dosage.<br />
– The high dosage should be a limit dose (1000 mg kg −1 ) or one inducing<br />
a minimal systemic effect on adults.<br />
– A short 2–4 week premating treatment period for both sexes <strong>of</strong> the<br />
parent (F0) generation.<br />
– A greater group size for the F0 generation to allow balanced selection <strong>of</strong><br />
the F1 generation.
294 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY<br />
Table 20.10 Detecting effects on males<br />
Notes:<br />
a Unknown=data not available or not examined<br />
A, organ weights, histopathology, serum chemistry.<br />
B, sperm analysis, count, motility, morphology.<br />
Data derived from an ongoing survey <strong>of</strong> 150 substances for which an effect on<br />
males has been claimed, mostly from human studies. To date 80 <strong>of</strong> the substances<br />
have been evaluated. Results indicate that use <strong>of</strong> a prolonged premating treatment<br />
period (>2 wks) has not been helpful for indicating effects on humans and is no<br />
better than use <strong>of</strong> a short premating period alone (
A.K.PALMER 295<br />
For example, a simple division <strong>of</strong> a two generation study will provide a<br />
developmental toxicity study in which pregnant females are treated from<br />
implantation through lactation or beyond and an F1 generation reared<br />
through to sexual maturity (Figure 20.8). The counterpart to this is a study<br />
<strong>of</strong> fertility in which both sexes are treated from about 2 weeks prior to<br />
Figure 20.6 A two generation study.
296 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY<br />
Figure 20.7 Reproductive toxicity—selecting studies.<br />
Figure 20.8 Pre- and post-natal (developmental toxicity) study.<br />
mating through to termination <strong>of</strong> males after a minimum <strong>of</strong> 4 weeks <strong>of</strong><br />
treatment overall (Figure 20.9). Treatment <strong>of</strong> females continues to<br />
implantation and they may be killed and examined at about day 13–15 <strong>of</strong><br />
pregnancy.<br />
Alternatively, treatment <strong>of</strong> females can be continued beyond closure <strong>of</strong><br />
the palate or even through pregnancy (Figure 20.4). Foetuses can be
Figure 20.9 Fertility<br />
delivered and examined for abnormalities according to procedures used in<br />
embryotoxicity studies. (Note that this study is almost identical to a<br />
modified OECD 421 study.) Combination with the ICH developmental<br />
toxicity study provides the equivalent <strong>of</strong> a two generation and<br />
embryotoxicity study.<br />
Interpretation <strong>of</strong> studies<br />
A.K.PALMER 297<br />
For regulatory agencies, one <strong>of</strong> the purposes <strong>of</strong> these tests is to gather<br />
information for labelling. By common consensus, a substance is labelled as<br />
a reproductive toxicant only if it induces effects at dosages below those<br />
causing systemic toxicity. Such labelling, however, can be very misleading<br />
especially with industrial chemicals. Even if the animal species can be<br />
shown to be a good surrogate for humans by means <strong>of</strong> kinetic and other<br />
studies, it is necessary to take into account the relationship between<br />
exposures causing effects and those likely to be encountered by humans.<br />
For example, a material can be labelled as a reproductive toxicant but<br />
present little or no real risk to humans because the effects are induced at<br />
exposures well in excess <strong>of</strong> those encountered by humans (Figure 20.10).<br />
Conversely, a substance not labelled as a reproductive toxicant could cause<br />
reproductive effects in humans if these are induced at exposures<br />
encountered by humans. In other words toxicity is relative, a matter <strong>of</strong><br />
dosage and situation. Labelling without reference to exposures is<br />
incomplete.
298 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY<br />
Figure 20.10 Interpretation/extrapolation <strong>of</strong> reproductive toxicology.<br />
Conclusions<br />
In conclusion I should mention that time constraints enforce superficial<br />
mention <strong>of</strong> many important aspects. I would like to emphasise that<br />
improved testing and evaluation for toxicity to reproduction can be<br />
achieved with methodology that exists within and without regulatory<br />
guidelines. This has been illustrated by the special cases presented at this<br />
meeting.<br />
There is no necessity to be restricted to specific guidelines, there never<br />
was. We should make use <strong>of</strong> any and all test methods available as<br />
appropriate for the substance being investigated. Whatever the type <strong>of</strong><br />
substance, testing involves looking for the same hazards. Having identified<br />
a hazard, methods <strong>of</strong> assessing risk, essentially, are the same (although<br />
PBPK models for reproductive toxicity would be more complex than those<br />
used for systemic toxicity).<br />
The methodology is available, what is required is the willingness and<br />
wisdom to use it effectively and efficiently (Palmer, 1993a, b). The goal<br />
should be to investigate a specific substance to the extent necessary, no<br />
more and no less. For this there needs to be a change in attitude by<br />
industry, agencies and academia. Therein is the greatest problem since<br />
‘Change is not made without inconvenience, even from worse to better’ and<br />
people are very unwilling to change their prejudices and habits.
References<br />
A.K.PALMER 299<br />
PALMER, A.K., 1993a, Identifying environmental factors harmful to reproduction,<br />
Environm. Hlth Perspect. Supplements, 101(2), 19–25.<br />
PALMER, A.K., 1993b, Introduction to (pre)screening methods, Reproduct.<br />
Toxicol., 7, 95–8.
PART SIX<br />
Toxicity <strong>of</strong> selected classes <strong>of</strong> industrial<br />
chemicals
21<br />
Special Points in the Toxicity Assessment <strong>of</strong><br />
Colorants (Dyes and Pigments)<br />
HERMANN M.BOLT<br />
Institut für Arbeitsphysiologie an der Universität Dortmund,<br />
Dortmund<br />
Introduction<br />
Colorants (dyes and pigments) are very important industrial chemicals. A<br />
special point in the toxicological assessment <strong>of</strong> such compounds is their<br />
bioavailability upon inhalation. From the technological point <strong>of</strong> view<br />
pigments are colorants which are insoluble whereas dyes are soluble in the<br />
application mixture.<br />
Biologically, the most relevant route <strong>of</strong> potential exposure <strong>of</strong> humans to<br />
colorants is by inhalation. If a pigment is biologically insoluble, it may<br />
finally be removed from the airways by clearance mechanisms. However, in<br />
practice the situation is much more complicated. For instance, chromates<br />
are technically important pigments which are well investigated. Biochemical<br />
and toxicological research has shown that the common toxicological<br />
principle <strong>of</strong> chromates which penetrate the cell membrane and, after<br />
intracellular transformation, exert genotoxic effects, is the chromate anion<br />
(CrO ). In terms <strong>of</strong> inhalatory carcinogenicity, the very water-soluble<br />
alkali chromates and the practically insoluble lead chromate have the<br />
lowest potency. Pigments <strong>of</strong> an intermediate solubility, e.g. calcium<br />
chromate, zinc chromate, strontium chromate, have a high carcinogenic<br />
potency on the respiratory tract. Local storage <strong>of</strong> chromate particles in the<br />
airways with a slow but continuous local release <strong>of</strong> CrO seems therefore<br />
to be an important factor in respiratory carcinogenesis induced by<br />
chromates. But also lead chromate, which is technically regarded as<br />
insoluble, is bioavailable to some extent; this is visualized by practical cases<br />
<strong>of</strong> occupational lead chromate exposure which display markedly elevated<br />
blood lead levels.<br />
The question <strong>of</strong> systemic bioavailability, upon inhalation, became <strong>of</strong><br />
recent regulatory importance for azo colorants based on carcinogenic<br />
aromatic amines. This problem has already been addressed in detail<br />
(Myslak and Bolt, 1988).
302 SPECIAL POINTS IN THE TOXICITY ASSESSMENT OF COLORANTS<br />
Table 21.1 Number <strong>of</strong> azo colorants based on cancerogenic aromatic amines (2napthylamine,<br />
benzidine and its derivatives) listed in Colour Index (3rd edn, 3rd<br />
Rev., 1987)<br />
The problem <strong>of</strong> carcinogenic azo colorants<br />
In the past, azo colorants based on benzidene, 3,3′-dichlorobenzidine, 3,3′dimethylbenzidine<br />
(o-tolidine), and 3,3'-dimethoxybenzidine (odianisidine)<br />
have been synthesized in large amounts and numbers,<br />
especially in the German chemical industry. The Colour Index (1987) lists<br />
a total number <strong>of</strong> more than 2000 azodyes, 452 <strong>of</strong> them being based on 2naphthylamine,<br />
benzidine, or benzidine derivatives (Table 21.1).<br />
Azo colorants have a number <strong>of</strong> properties that have made them<br />
invaluable for dyeing a wide variety <strong>of</strong> materials, including natural and<br />
artificial fibres, plastics, resins, textiles, leather, paper, glass, ceramics,<br />
cement, inks, printing inks, chalks, crayons and carbon papers, as well as<br />
cosmetics, food and beverages. Interesting with respect to potential<br />
exposure <strong>of</strong> painters is the use <strong>of</strong> azo colorants in the coloring <strong>of</strong> oil-,<br />
resins-, emulsion-, lime-, and other aqueousbased paints, distempers,<br />
transparent laquers, spirit and oil wood stains, and varnishes (Colour<br />
Index, 1987). In all these fields, particularly benzidine-based azo colorants<br />
have found widespread use (Gregory, 1984).<br />
In the UK, the Carcinogenic Substances Regulation led in 1967 to<br />
discontinuation <strong>of</strong> the use <strong>of</strong> benzidine in the production <strong>of</strong> azo colorants<br />
(Martin and Kennelly, 1985). The US government in 1974 promulgated<br />
regulations to control benzidine at the workplace (Gregory, 1984).<br />
Nevertheless, in the period <strong>of</strong> 1972–4, more than 150000 persons in the<br />
USA were potentially occupationally exposed to benzidine-based colorants<br />
(Gregory, 1984); in 1978, approximately 1.7 million US pounds <strong>of</strong><br />
benzidine-based azo colorants were manufactured, and a further 1.6<br />
million pounds were imported into the USA (Lynn et al., 1980).<br />
In Germany, over 30 different benzidine-based azo colorants were<br />
manufactured in the early 1960s. The manufacture <strong>of</strong> these colorants was<br />
stopped in 1971, with the exception <strong>of</strong> one dye (Direct Black 4; C.I. No.<br />
30245); the manufacture <strong>of</strong> the latter was continued until 1973. Azo
H.M.BOLT 303<br />
colorants based on carcinogenic congeners <strong>of</strong> benzidine (e.g. 3,3′dimethoxybenzidine;<br />
3,3′-dimethylbenzidine) are most likely still being<br />
manufactured in some countries. The case <strong>of</strong> pigments based on 3,3′dichlorobenzidine<br />
is discussed below.<br />
Azo colorants are biologically active through their metabolites.<br />
Azoreduction <strong>of</strong> these compounds occurs in vivo (Radomski and<br />
Mellinger, 1962; Rinde and Troll, 1975; Robens et al., 1980) by an<br />
enzyme-mediated reaction. Azoreductases are found in mammalian tissues,<br />
particularly in liver (Fouts et al., 1957; Walker, 1970; Martin and Kennelly,<br />
1981; Kennelly et al., 1982) and also in gut bacteria (Yoshida and<br />
Miyakawa, 1973; Chung et al., 1978; Hartmann et al., 1978; Cerniglia et<br />
al., 1982; Bos et al., 1986). The result <strong>of</strong> this azoreduction is the release <strong>of</strong><br />
the (carcinogenic) aromatic amine from the colorant (Martin and Kennelly,<br />
1985). Studies performed on exposed workers have demonstrated that the<br />
azoreduction <strong>of</strong> benzidine-based colorants occurs in man (Genin, 1977;<br />
Boeninger, 1978; Lowry et al., 1980; Meal et al., 1981; Dewan et al.,<br />
1988). Studies <strong>of</strong> Lynn et al., (1980) and Bowman et al. (1983) have<br />
demonstrated that the metabolic conversion <strong>of</strong> benzidine-, 3,3′dimethylbenzidine-<br />
and 3,3′-dimethoxybenzidine-based colorants to their<br />
(carcinogenic) amine precursors in vivo is a general phenomenon that must<br />
be expected for each member <strong>of</strong> this class <strong>of</strong> chemicals.<br />
However, in contrast to water-soluble dyes, the question <strong>of</strong> biological<br />
azoreduction <strong>of</strong> (practically insoluble) pigments was a matter <strong>of</strong> discussion<br />
in the recent years. One study has claimed the presence <strong>of</strong> 3,3'dichlorobenzidine<br />
in the urine both <strong>of</strong> experimental animals fed with<br />
Pigment Yellow 12 and <strong>of</strong> exposed workers (Akiyama, 1970). However,<br />
other experimental studies, using more modern analytical tools, did not<br />
confirm these results (DHEW, 1978; Leuschner, 1978; Mondino et al.,<br />
1978; Nony et al., 1980).<br />
Several epidemiological studies have demonstrated that the use <strong>of</strong> the<br />
benzidine-based dyes has caused bladder cancer in humans. In a Japanese<br />
study, the risk <strong>of</strong> bladder cancer among dye applicators (kimono painters)<br />
was 6.8 times the expected rate (Yoshida et al., 1971). In a British study,<br />
workers performing the dyeing <strong>of</strong> textiles (and not exposed to benzidine<br />
itself) had a higher risk <strong>of</strong> bladder cancer (RR=3.4) than expected<br />
(Anthony, 1974). In a USSR study, an increased incidence <strong>of</strong> bladder<br />
cancer was found in workers who dried or ground benzidine-based dyes<br />
(Genin, 1977).<br />
In our own study on bladder cancer in painters (Myslak et al., 1991), the<br />
time <strong>of</strong> first exposure (painters with bladder tumors) dated mostly back to<br />
the first half <strong>of</strong> the century. Two factors may have been relevant: (1) at<br />
that time, a large number <strong>of</strong> benzidine-based azodyes was in manufacture,<br />
especially in Germany; (2) during that time it was usual for painters in<br />
Germany to prepare their paints themselves. This work included grinding
304 SPECIAL POINTS IN THE TOXICITY ASSESSMENT OF COLORANTS<br />
and mixing <strong>of</strong> the dyes and preparation <strong>of</strong> the coloring mixture by addition<br />
<strong>of</strong> solvents. All painters we had interviewed reported that this type <strong>of</strong> work<br />
had been regularly associated with considerable occurrence <strong>of</strong> dye dust in<br />
the atmosphere, up to the end <strong>of</strong> the 1950s.<br />
The very long latency period may explain why an enhanced risk <strong>of</strong> bladder<br />
cancer in German painters (due to previous exposure to azo dyes) is<br />
observed even today. Similar arguments have also been put forward for<br />
other occupational groups associated with an increased risk <strong>of</strong> bladder<br />
cancer, and where a causal connection with benzidine-based azo dyes had<br />
been proven or suggested, e.g. for textile dyers (Jenkins, 1978), leather<br />
dyers and shoeworkers (Decouflé, 1979), hairdressers (Guberan et al.,<br />
1985), and tailors (Anthony and Thomas, 1970). The results <strong>of</strong> our own<br />
survey <strong>of</strong> painters are very probably not relevant for the present working<br />
conditions in Germany and other highly industrialized countries, because<br />
<strong>of</strong> different materials, working methods, and hygienic standards introduced<br />
in recent years. They are, however, quite relevant for matters <strong>of</strong><br />
compensation <strong>of</strong> persons who are now diseased.<br />
Regulatory aspects (FRG)<br />
The arguments described above have led the German Commission for<br />
Investigation <strong>of</strong> Health Hazards <strong>of</strong> Chemical <strong>Compounds</strong> in the Work<br />
Area (MAK-Commission) to include the following chapter in the MAKlist,<br />
since 1988 (DFG, 1988):<br />
Azo colorants are characterized by the azo group -N=N-. They are<br />
made by the coupling <strong>of</strong> singly and multiply diazotized aryl amines.<br />
Of particular toxicological importance are colorants from double<br />
diazotized benzidine and from benzidine derivatives (3,3′dimethylbenzidine,<br />
3,3′-dimethoxybenzidine, 3,3′-dichlorobenzidine).<br />
In addition, aminoazo-benzene, aminonaphthalene and monocyclic<br />
aromatic amines are encountered. Reductive fission <strong>of</strong> the azo group,<br />
either by intestinal bacteria or by azo reductases <strong>of</strong> the liver and<br />
extrahepatic tissues, can cause these compounds to be released. Such<br />
breakdown products have been detected in animal experiments as<br />
well as in man (urine). Mutagenicity, which has been observed with<br />
numerous azo colorants in in-vitro test systems, and the<br />
carcinogenicity in animal experiments are attributed to the release <strong>of</strong><br />
amines and their subsequent metabolic activation. There are now<br />
epidemiological indications that occupational exposure to benzidinebased<br />
azo colorants can increase the incidence <strong>of</strong> bladder carcinomas.<br />
Thus, all azo colorants whose metabolism can liberate a carcinogenic<br />
aryl amine are suspected <strong>of</strong> having carcinogenic potential. Due to the<br />
large number <strong>of</strong> such dyes (several hundred) it seems neither possible
nor justifiable to substantiate this suspicion in each individual case by<br />
means <strong>of</strong> animal experimentation according to customary<br />
classification criteria. Instead, scientifically justifiable models have to<br />
be relied on. Therefore, as a preventive measure to avoid putting<br />
exposed persons at risk, it is recommended that the substances be dealt<br />
with as if they were classified in the same categories as the<br />
corresponding carcinogenic or suspected carcinogenic amines (A1, A2,<br />
B)<br />
If there are indications that the colorant itself (e.g. a pigment) or<br />
any carcinogenic breakdown products are not biologically available,<br />
the exclusion <strong>of</strong> risk should be experimentally proven or<br />
substantiated by biomonitoring. Suitable animal experiments can also<br />
rule out suspicion <strong>of</strong> carcinogenic potential.<br />
On the basis <strong>of</strong> this general view, which had been endorsed by the German<br />
Ministry <strong>of</strong> Labor (TGS 900, Bundesarbeitsblatt 1/1990, p. 63), the<br />
identification <strong>of</strong> the aromatic amine component <strong>of</strong> azo colorants is <strong>of</strong> key<br />
importance. A suitable compilation <strong>of</strong> azo colorants, according to their<br />
aromatic amine components, has been published by Myslak (1990).<br />
Azo pigments<br />
H.M.BOLT 305<br />
The postulate <strong>of</strong> further research on the bioavailability and/or<br />
carcinogenicity <strong>of</strong> azo pigments, especially those based on 3,3′dichlorobenzidine<br />
(v.s.), has focused interest on this particular problem.<br />
Azo pigments based on 3,3′-dichlorobenzidine (e.g. Pigment Yellow 12,<br />
Pigment Yellow 13, Pigment Yellow 14) have been orally administered to<br />
rats, hamsters, rabbits and monkeys, at doses up to 400 mg pigment kg −1<br />
b.w. (Leuschner, 1978; Mondino et al., 1978; DHEW, 1978; Nony et al.,<br />
1980; Decad et al., 1983; Sagelsdorff et al., 1990; Hoechst AG,<br />
unpublished data). These authors could not find 3,3'-dichlorobenzidine in<br />
the urine <strong>of</strong> animals treated with 3,3'-dichlorobenzidine pigments. Decad<br />
et al. (1983) demonstrated that 14 C-labelled Pigment Yellow 12, orally<br />
administered to rats, was completely excreted in the faeces. On the basis <strong>of</strong><br />
these investigations <strong>of</strong> disposition <strong>of</strong> 3,3′-dichlorobenzidine-based<br />
pigments, it is clear why none <strong>of</strong> the long-term animal carcinogenicity<br />
studies performed so far (Leuschner, 1978; DHWE 1978; ICI, unpublished<br />
data) has demonstrated a carcinogenic effect <strong>of</strong> a diaryl pigment.<br />
It therefore appears that the aromatic amine components from azo<br />
pigments are practically not bioavailable, as demonstrated for several<br />
pigments on the basis <strong>of</strong> 3,3′-dichlorobenzidine (ETAD, 1990; see also Bolt<br />
and Golka, 1993). Hence, it is now very unlikely that occupational<br />
exposure to insoluble azo pigments would be associated with a substantial<br />
risk <strong>of</strong> (bladder) cancer in man.
306 SPECIAL POINTS IN THE TOXICITY ASSESSMENT OF COLORANTS<br />
References<br />
AKIYAMA, T., 1970, The investigation on the manufacturing plant <strong>of</strong> organic<br />
pigment, Jikei Med. J., 17, 1–9.<br />
ANTHONY, H.M., 1974, <strong>Industrial</strong> exposure in patients with carcinoma <strong>of</strong> the<br />
bladder, J. Soc. Occup. Med., 24, 110–16.<br />
ANTHONY, H.M. and THOMAS, G.M., 1970, Tumors <strong>of</strong> the urinary bladder:<br />
An analysis <strong>of</strong> the occupations <strong>of</strong> 1030 patients in Leeds, England, J. Nat.<br />
Cancer Inst., 45, 879–95.<br />
BOENINGER, H., 1978, An investigation <strong>of</strong> the metabolic reduction <strong>of</strong> benzidine<br />
azo dyes to benzidine and its metabolites and their possible relationship to<br />
carcinoma <strong>of</strong> the bladder in man. Unpublished data (cited by Gregory, 1984).<br />
BOLT, H. M and GOLKA, K., 1993, Zur früheren Exposition von Malern<br />
gegenüber Az<strong>of</strong>arbst<strong>of</strong>fen, Arbeitsmed., Sozialmed., Umweltmed., 28, 417–21.<br />
Bos, R.P., VAN DER KRIEKEN, W., SMEIJSTERS, L., KOOPMAN, J.P.,<br />
DEJONGE, H.R., THEUWS, J.L.G. and HENDERSON, P.T., 1986, Internal<br />
exposure <strong>of</strong> rats to benzidine derived from orally administered benzidine-based<br />
dyes after intestinal azo reduction, <strong>Toxicology</strong>, 40, 207–13.<br />
BOWMAN, M.C., NONY, C.R., BILLEDEAU, S.M., MARTIN, J.L. and<br />
THOMPSON, H.C., 1983, Metabolism <strong>of</strong> nine benzidine-congener-based azo<br />
dyes in rats based on gas chromatographic assays <strong>of</strong> the urine for potentially<br />
carcinogenic metabolites, J. Anal. Toxicol. 7, 55–60.<br />
CERNIGLIA, C.E., FREEMAN, J.P., FRANKLIN, W. and PACK, L.D., 1982,<br />
Metabolism <strong>of</strong> azo dyes derived from benzidine, 3,3'-dimethylbenzidine and 3,<br />
3'-dimethoxybenzidine to potentially carcinogenic aromatic amines by<br />
intestinal bacteria, Carcinogenesis, 3, 1255–60.<br />
CHUNG, K.T., FULK, G.E. and EGAN, M., 1978, Reduction <strong>of</strong> azo dyes by<br />
intestinal anaerobes, Appl. Environ. Microbiol. 35, 55–62.<br />
Colour Index, 3rd edn., 3rd rev., 1987, Bradford: Society <strong>of</strong> Dyers and Colourists.<br />
Vols. 1–8.<br />
DECAD, G.M., SNYDER, C.D. and MITONA, C., 1983, Fate <strong>of</strong> water-insoluble<br />
and water-soluble dichlorobenzidine-based pigments, J. Toxicol. Environm.<br />
Hlth, 11, 455–65.<br />
DECOUFLÉ, P., 1979, Cancer risk associated with employment in the leather and<br />
leather products industry, Arch. Environm. Hlth, 34, 33–7.<br />
DFG (Deutsche Forschungsgemeinschaft), 1988, List <strong>of</strong> MAK and BAT Values<br />
1988, Weinheim: VCH Publishers.<br />
DEWAN, A., JANI, J.P., PATEL, J.S., GANDHI, D.N., VARIYA, M.R. and<br />
GHODSARA, N.B., 1988, Benzidine and its acetylated metabolites in the urine<br />
<strong>of</strong> workers exposed to Direct Black 38, Arch. Environm. Hlth, 43, 269–72.<br />
DHEW, 1978, Bioassay <strong>of</strong> diarylanilide yellow for possible carcinogenicity, DHEW<br />
Publication No. (NIH) 78–830, US Dept <strong>of</strong> Health, Education and Welfare,<br />
Public Health Service, National Cancer Institute, Carcinogens Testing Program.<br />
ETAD (Ecological and Toxicological Association <strong>of</strong> the Dyestuffs Manufacturing<br />
Industry, 1990, Zum kanzerogenen Potential von Diaryl-Azopigmenten auf<br />
Basis von 3,3′-Dichlorbenzidin, ETAD-Bericht T 2028-BB (D), ETAD,<br />
CH-4005, Basel 5.
H.M.BOLT 307<br />
FOUTS, J.R., KAMM, J.J. and BRODIE, B.B., 1957, Enzymatic reduction <strong>of</strong> prontosil<br />
and other azo dyes, J. Pharmacol, Exp. Ther., 110, 291–300.<br />
GENIN, W.A., 1977, Formation <strong>of</strong> clastogenic diphenylamine derivates as a result<br />
<strong>of</strong> the metabolism <strong>of</strong> direct azo dyes, Vopr. Oncol, 23, 50–2 (in Russian).<br />
GREGORY, A., 1984, The carcinogenic potential <strong>of</strong> benzidine-based dyes, J.<br />
Environm. Pathol. Toxicol. Oncol, 5, 243–59.<br />
GUBERAN, R., RAYMOND, L. and SWEETNAM, P.M., 1985, Increased risk for<br />
male bladder cancer among a cohort <strong>of</strong> male and female hairdressers from<br />
Geneva, Int. J. Epidemiol, 14, 549–54.<br />
HARTMANN, C.P., FULK, G.E. and ANDREWS, A.W., 1978, Azo reduction <strong>of</strong><br />
trypan blue to a known carcinogen by a cellfree extract <strong>of</strong> a human<br />
intestinal anaerobe, Mutat. Res., 58, 125–32.<br />
JENKINS, C.L., 1978, Textile dyes are potential hazards, J. Environm. Hlth, 40,<br />
279– 84.<br />
KENNELLY, J.C., HERTZOG, P.J. and MARTIN, C.N., 1982, The release <strong>of</strong> 4,4′diaminobiphenyls<br />
from azo dyes in the rat, Carcinogenesis, 3, 947–51.<br />
LEUSCHNER, F., 1978, Carcinogenicity studies on different diarylide yellow<br />
pigments in mice and rats, Toxicol. Lett., 2, 253–60.<br />
LOWRY, L.K., TOLOS, W.P., BOENINGER, M.F., NONY, C.R. and<br />
BOWMAN, M.C., 1980, Chemical monitoring <strong>of</strong> urine from workers<br />
potentially exposed to benzidine-derived azo dyes, Toxicol. Lett., 7, 29–36.<br />
LYNN, R.K., DONIELSON, D.W., ILIAS, A.M., KENNISH, J.M., WONG, K. and<br />
MATHEWS, H.B., 1980, Metabolism <strong>of</strong> bisazobiphenyl dyes derived from<br />
benzidine, 3,3′-dimethylbenzidine or 3,3′-dimethoxybenzidine to carcinogenic<br />
aromatic amines in the dog and rat, Toxicol Appl. Pharmacol, 56, 248–58.<br />
MARTIN, C.N. and KENNELLY, J.C., 1981, Rat liver microsomal azoreductase<br />
activity on four azo dyes derived from benzidine, 3,3′-dimethylbenzidine or 3,3′dimethoxybenzidine,<br />
Carcinogenesis, 2, 307–12.<br />
MARTIN, C.N. and KENNELLY, J.C., 1985, Metabolism, mutagenicity and DNA<br />
biding <strong>of</strong> biphenyl-based azo dyes, Drug Metab. Rev., 16, 89–117.<br />
MEAL, P.F., COCKER, J., WILSON, H.K. and GILMOUR, J.M., 1981, Search for<br />
benzidine and its metabolites in urine <strong>of</strong> workers weighing benzidine-derived<br />
dyes, Br. J. Ind. Med., 38, 191–3.<br />
MONDINO, A., ACHARI, R., DUBINI, M., MARCHISIO, M.A., SILVESTRI, S.<br />
and ZANOLO, G., 1978, Absence <strong>of</strong> dichlorobenzidine in the urine <strong>of</strong> rats,<br />
rabbits and monkeys treated with C.I. Pigment Yellow 13, Med. Lav., 69, 693–<br />
7.<br />
MYSLAK, Z.W., 1990, Az<strong>of</strong>arbmittel auf der Basis krebserzeugender und<br />
verdächtiger aromatischer Amine. Identification, Verwendungsbereiche,<br />
Herstellungszeiträume. Schriftenreihe der Bundesanstalt für Arbeitsschutz, GA<br />
35, Bremerhaven: Wissenschaftsverlag NW.<br />
MYSLAK, Z.W. and BOLT, H.M., 1988, Berufliche Exposition gegenüber<br />
Az<strong>of</strong>arbst<strong>of</strong>fen und Harnblasenkarzinom-Risiko, Zbl. Arbeitsmed., 10, 310–<br />
21.<br />
MYSLAK, Z.W., BOLT, H.M. and BROCKMANN, W., 1991, Tumors <strong>of</strong> the<br />
urinary bladder in painters: a case-control study, Am. J. Ind. Med., 19, 705–13.<br />
NONY, C.R., BOWMAN, M.C., CAIRNS, T., LOWRY, L.K. and TOLOS, W.P.,<br />
1980, Metabolism studies <strong>of</strong> an azo dye and pigment in the hamster based on
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analysis <strong>of</strong> the urine for potentially carcinogenic aromatic amine metabolites.<br />
J. Anal. Toxicol, 4, 132–40.<br />
RADOMSKI, J.L. and MELLINGER, T.J., 1962, The absorption, fate and<br />
excretion in rats <strong>of</strong> the water-soluble azo dyes, FD&C Red No. 2. FD&C Red<br />
No. 4 and FD&C Yellow No. 6, J. Pharmacol. Exp. Ther., 136, 259–66.<br />
RINDE, E. and TROLL, W., 1975, Metabolic reduction <strong>of</strong> benzidine azo dyes to<br />
benzidine in the Rhesus monkey, J. Nat. Cancer Inst., 55, 181–2.<br />
ROBENS, J.F., DILL, G.S., WARD, J.M., JOINER, J.R., GRIESEMER, R.A. and<br />
DOUGLAS, J.F., 1980, Thirteen-week subchronic toxicity studies <strong>of</strong> Direct<br />
Blue 6, Direct Black 38 and Direct Brown 95 dyes, Toxicol. Appl. Pharmacol.,<br />
54, 431–42.<br />
SAGELSDORFF, P., JOPPICH-KUHN, M. and JOPPICH, M., 1990,<br />
Biomonitoring for the bioavailability <strong>of</strong> dichlorobenzidine from<br />
dichlorobenzidine-based dyes, J. Cancer Res. Clin. Oncol, 116, 79 (abstract).<br />
WALKER, R., 1970, The metabolism <strong>of</strong> azo compounds, a review <strong>of</strong> the literature,<br />
Food Cosmet. Toxicol, 8, 659–76.<br />
YOSHIDA, O., HARADA, T., MIYAKAWA, M. and KATO, T., 1971, Bladder<br />
cancer among dyers in the Kyoto area, Igaku Ayumi, 79, 421–2 (in Japanese).<br />
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aspects, in Nakahara, W., Hirayama, T., Nishioka, K. and Sugano, H. (Eds)<br />
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Symp. <strong>of</strong> Princess Takamatsu Cancer Research Fund, pp. 31–9, Baltimore:<br />
University Park Press.
22<br />
<strong>Toxicology</strong> <strong>of</strong> Textile Chemicals<br />
DIETER SEDLAK<br />
EnviroTex GmbH, Augsburg<br />
The former main aspects <strong>of</strong> textiles like fashion or usefulness seem to be<br />
pushed back by a new phenomenon, the ecological and toxicological aspect<br />
<strong>of</strong> textiles. Although everybody is talking about textiles in this context,<br />
everybody means the textile chemicals (and dye-stuffs) on the textile. With<br />
respect to this situation we have to consider new developments in Germany<br />
(Figure 22.1).<br />
For two years now textile finishing plants have to be approved within<br />
the German federal immission control act. This means that all sorts <strong>of</strong><br />
immissions to the working place or the surroundings <strong>of</strong> the plant must be<br />
defined in quantity and quality. In the meantime so-called emission factors<br />
have been developed. But this project is not yet finished. Even for insiders<br />
it was surprising how many textile chemicals show unexpected behaviour<br />
during their application. The reasons are:<br />
– unknown byproducts or impurities,<br />
– unknown reactivities between components,<br />
– unknown interactions with substrate,<br />
– unknown dependencies <strong>of</strong> process parameters.<br />
Today producers and users realise that the inherent toxicological properties<br />
<strong>of</strong> many textile chemicals evaluated within a typical safety data sheet<br />
represent only a part <strong>of</strong> the whole knowledge you need for safe handling<br />
and processing.<br />
A further interesting development undoubtedly is the discussion around<br />
the ecolabelling <strong>of</strong> textiles. There definitely is some danger so that different<br />
labels based on different commercial interests should be evaluated. A<br />
positive step could be the fact that MST and Ecotex have joined. But other<br />
societies like TÜV or GSF create their own labels including statements that<br />
they use the better (right) label criteria. This leads quite clearly to total<br />
confusion for the consumer. However, what the textile industry needs<br />
today is confidence. The developments discussed above will not help. From<br />
a technical point <strong>of</strong> view we have the same problem as discussed with<br />
respect to the immissions. Too little is known about the real properties and
310 TOXICOLOGY OF TEXTILE CHEMICALS<br />
Figure 22.1 New developments in Germany regarding toxicology <strong>of</strong> textile<br />
chemicals.<br />
behaviour <strong>of</strong> textile chemicals to define the absolute label criteria. These<br />
uncertainties can be solved only by cooperation <strong>of</strong> the different groups and<br />
not by aggressive competition. How can we handle both developments in a<br />
proper way?<br />
At first we have to be clear about the interactions <strong>of</strong> both problem fields.<br />
Let us start with the textile chemical delivered to a finishing plant<br />
(Figures 22.2 and 22.3). Following the directives for dangerous substances<br />
or safety data sheets the product is exactly labelled and its toxicological
Figure 22.2 Possible levels to discuss the toxicology <strong>of</strong> textile chemicals.<br />
D.SEDLAK 311<br />
properties are well described in the SDS. This is only as good as the<br />
information given to the product safety manager responsible for the<br />
possible ways <strong>of</strong> handling and processing. However, in most cases textile<br />
chemicals can be described as harmless.<br />
Nearly all show an acute oral toxicity greater than 2000 mg kg −1<br />
although they may contain toxic substances in diluted form. Only a few <strong>of</strong><br />
them possess an irritant character or even CMT properties. Nevertheless,<br />
many <strong>of</strong> them have impurities with these properties which are given more<br />
and more, even in low concentrations. This is a very important process
312 TOXICOLOGY OF TEXTILE CHEMICALS<br />
Figure 22.3 Toxicological pr<strong>of</strong>ile <strong>of</strong> textile chemicals.<br />
because these low concentrations may also lead to high vapour<br />
concentrations at the workplace if these substances are volatile. This is<br />
mostly the case.<br />
For example, the carcinogenic substance acrylonitrile has a product label<br />
only giving a concentration <strong>of</strong> 0.2 percent and more in the corresponding<br />
polymer dispersion. Even 0.02 percent may lead to workplace<br />
concentrations a hundred fold higher than the TLV allows. Product data<br />
sheets with these ‘properties’ still exist with no additional information<br />
because it is not needed!<br />
Another problem is that nearly all textile chemicals are exposed to<br />
temperatures between 100°C and 230°C during application which leads to<br />
many only poorly defined substances which may all be set free in the<br />
workplace and the surroundings. Here we are also confronted with a<br />
typical juristic phenomenon. Many suppliers do not take any responsibility<br />
regarding these questions if the user defines his process parameters himself,<br />
especially when using recipe components from different suppliers. But<br />
everybody knows that the user has not the means to evaluate the<br />
toxicological pr<strong>of</strong>ile <strong>of</strong> his chemicals mixture and process. This lack <strong>of</strong><br />
responsibility should be clarified.<br />
Now let us observe the result <strong>of</strong> such chemical treatment <strong>of</strong> textiles.<br />
Normally you will only discuss the summarised toxicological pr<strong>of</strong>iles <strong>of</strong> the<br />
used chemicals in an additive way. This means that the combination <strong>of</strong><br />
different products—all with no acute toxicity—will also show no acute
Figure 22.4 Typical composition <strong>of</strong> a flame retardant recipe.<br />
D.SEDLAK 313<br />
toxicity. Furthermore the concentration <strong>of</strong> the active ingredients in the new<br />
matrix textile is much lower than in the matrix water most textile<br />
chemicals are based on. This strategy seems to be appropriate. On the<br />
other hand take the irritant property <strong>of</strong>, for example, fatty amine based<br />
emulsifier. This property is lost during formulation <strong>of</strong> the emulsifier in the<br />
textile chemical by homogeneous dilution. The average concentration after<br />
application to the textile is surely lower. But who has information about<br />
the actual form <strong>of</strong> this fatty amine in the textile. Is it distributed<br />
homogeneously, is it more located at the surface, does it interact with other<br />
substances and even lose its irritant character? Many questions seem to be<br />
unanswered.<br />
Both complex questions—the toxicology <strong>of</strong> the handling and processing<br />
<strong>of</strong> textile chemicals and the toxicology <strong>of</strong> the result <strong>of</strong> the processing on<br />
textiles— will be discussed by means <strong>of</strong> an admittedly drastic example, a<br />
flame retardant process. This example is rather representative because an
314 TOXICOLOGY OF TEXTILE CHEMICALS<br />
Figure 22.5 Total composition <strong>of</strong> a flame retardant recipe.<br />
EEC-directive is ‘under construction’ just now which demands the finishing<br />
<strong>of</strong> all upholstery covers with flame retardant chemicals.<br />
The composition <strong>of</strong> this recipe, shown in Figure 22.4, doesn’t seem to be<br />
too complex. No special toxicological properties are apparent. The<br />
corrosive character <strong>of</strong> the phosporic acid vanishes during the application
D.SEDLAK 315<br />
Figure 22.6 Release <strong>of</strong> chemical substances during/after a flame retardant process.<br />
by neutralizing the finished textile with alkalis. Complex reactions <strong>of</strong> the<br />
different components are expected which are supposed to form not well<br />
defined polycondensates fixed to the textile fibre. On this basis there is no<br />
apparent reason to think <strong>of</strong> any toxicological side effects.<br />
In Figure 22.5, however, you will get a good idea about the real<br />
situation. But this detailed composition should not be deceptive about the<br />
fact that this is the ‘wanted’ technical composition verified by a few<br />
analytical data. In this case about 500 actual substances can be expected.<br />
Even if we should have available all the necessary toxicological data for the<br />
components it would not help, because there is poor information about the<br />
result <strong>of</strong> the finishing process itself. This is the only reality.<br />
A first step in the right direction would be the analysis <strong>of</strong> the process<br />
<strong>of</strong>fgas. The result reflects approximately the activities on the textile during
316 TOXICOLOGY OF TEXTILE CHEMICALS<br />
Figure 22.7 Needs for the textile toxicology assessment.<br />
processing. Figure 22.6 shows in which way this knowledge can be used to<br />
perform a proper risk assessment based on defined substances. However, in<br />
the case <strong>of</strong> evaluation <strong>of</strong> the skin toxicity we are confronted with a lack <strong>of</strong><br />
data concerning the bio-availability <strong>of</strong> the named substances. Not before<br />
having cleared these additional questions can the toxicologist seriously<br />
start his routine activities, the risk assessment. Anyway this seems to be<br />
much easier than to discover the chemical basis for this assessment.<br />
With respect to the toxicology <strong>of</strong> textile chemicals there is a lot to do<br />
mostly in the field <strong>of</strong> understanding the chemicals and processes used.<br />
After this we recommend the development <strong>of</strong> models for evaluating the<br />
bioavailability <strong>of</strong> substances and test kits to characterize the toxicological<br />
properties <strong>of</strong> textiles as a whole. Last but not least we hope for much<br />
better communication between the industries involved (Figure 22.7).
23<br />
Antioxidants and Light Stabilisers : Toxic Effects<br />
<strong>of</strong> 3,5-Di-alkyl-4hydroxyphenyl Propionic Acid<br />
Derivatives in the Rat and their Relevance for<br />
Human Safety Evaluation<br />
HELMUT THOMAS, PETER DOLLENMEIER, ELKE<br />
PERSOHN, HANSJÖRG WEIDELI and FELIX WAECHTER<br />
Ciba-Geigy Limited, Basle<br />
Introduction<br />
Antioxidants and light stabilisers are important additives for a wide range<br />
<strong>of</strong> plastic materials used for industrial as well as for medicinal and food<br />
packaging purposes. Technical efficiency requires that these compounds are<br />
mobile within the polymer network. This implies that humans may be<br />
exposed to such compounds not only during the manufacturing process but<br />
also in the course <strong>of</strong> the decay <strong>of</strong> polymers, by direct contact with<br />
substances that have migrated to the surface <strong>of</strong> plastic materials or by<br />
ingestion <strong>of</strong> diffusion-contaminated food. It is <strong>of</strong> considerable interest,<br />
therefore, to be aware <strong>of</strong> the toxicological properties <strong>of</strong> these compounds<br />
and the relevance <strong>of</strong> these properties for the safety assessment in humans.<br />
Sterically hindered phenolic antioxidants<br />
Phenolic antioxidants have been widely used as food preservatives and<br />
their almost classical representatives, 2,6-di-tert-butyl-4-methyl phenol<br />
(BHT) and 2-tert-butyl-4-methoxyphenol (BHA) have been characterised in<br />
extenso with respect to their biochemical and toxicological properties<br />
(Søndergaard and Olsen, 1982; Conning and Phillips, 1986; Ito et al.,<br />
1986a; Williams et al., 1990a,b; Verhagen et al., 1991). Being nongenotoxic,<br />
these compounds were found upon oral administration to<br />
rodents to cause slightly increased liver weight, to induce mainly epoxide<br />
hydrolase and phase II drug metabolising enzymes and to exert some anticarcinogenic<br />
effect (Benson et al., 1979; Choe et al., 1984; Gregus and<br />
Klaassen, 1988; Prochaska and Talalay, 1988; Perchellet and Perchellet,<br />
1989; Rodrigues et al., 1991). Pulmonary toxicity and carcinogenicity<br />
particularly <strong>of</strong> BHT in the mouse and formation <strong>of</strong> forestomach carcinoma<br />
and papilloma in the rat and Syrian hamster, respectively, have been<br />
reported and are considered to be largely species-specific effects (Abraham<br />
et al., 1986; Anon, 1986; Ito et al., 1986a, b; Verhagen et al., 1989).<br />
<strong>Industrial</strong> phenolic antioxidants as used in the polymer technology, by
318 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
contrast, although related to BHT, require higher molecular weights and<br />
hydrophobicity in order to retain them in the polymer matrix. This can be<br />
achieved, for example, by introducing alkyl chains or (esterified)<br />
carboxyalkyl moieties in the para-position to the phenolic hydroxyl group<br />
with variations in the aliphatic substitution pattern <strong>of</strong> the 2- and 6positions.<br />
With these modifications the questions arose, whether or not the<br />
toxicology <strong>of</strong> the chemically modified <strong>of</strong>fspring would still be related to<br />
that <strong>of</strong> the phenolic core <strong>of</strong> the BHT ancestor or completely different,<br />
although a common metabolic degradation pathway to structural entities<br />
resembling BHT could be anticipated. Clarification <strong>of</strong> this question was<br />
expected to contribute to the understanding <strong>of</strong> the toxicology <strong>of</strong> an entire<br />
class <strong>of</strong> phenolic antioxidants. A number <strong>of</strong> differentially esterified 4hydroxy-3,<br />
5-dialkyl-phenylpropionic acid derivatives were subsequently<br />
subjected to subchronic toxicity testing in rats with the result that the<br />
effects encountered were largely dependent on the alcohol moiety used for<br />
esterification and the size <strong>of</strong> the alkyl-substituent in the 3- and 5-positions:<br />
most frequently, increased liver weights were encountered in compounds<br />
with bulky 3,5-substituents such as tert-butyl. In some instances, however,<br />
initial hepatomegaly was followed after several weeks or months <strong>of</strong><br />
treatment by increases in thyroid weights and a proliferation <strong>of</strong> the thyroid<br />
follicular epithelium. The mechanism leading to the latter effect has been<br />
investigated in detail using the di-ester <strong>of</strong> 3-tert-buty1–4-hydroxy-5-methylphenylpropionic<br />
acid with ethylene glycol, Compound B, as a model<br />
compound (Table 23.1).<br />
Blood kinetics and blood metabolites<br />
Compound B is a symmetrical di-ester compound (Table 23.1). When<br />
administered at a single oral dose <strong>of</strong> about 10 mg kg −1 body weight to male<br />
rats, 14 C-phenyl-labelled Compound B was readily adsorbed, and maximal<br />
blood radioactivities were reached after 1 h. Thereafter, blood radioactivity<br />
declined rapidly and only minute amounts were detected 48 h after<br />
treatment. At any time point investigated Compound A, the free carboxylic<br />
acid <strong>of</strong> Compound B, was the dominating blood metabolite, whereas the<br />
parent compound constituted a minor component only (Table 23.2). These<br />
findings are indicative <strong>of</strong> an efficient first pass hydrolysis and suggest that<br />
the carboxylic acid metabolite, Compound A, might be responsible for the<br />
toxicological pr<strong>of</strong>ile <strong>of</strong> this antioxidant in the rat.<br />
Liver enzyme induction<br />
Compound B was administered in the feed to male rats at dose levels <strong>of</strong> 50,<br />
150, 500 and 1000 ppm. After treatment for 14 days, absolute liver<br />
weights were dose-dependently increased. Biochemically, this
Table 23.1 3.5-Substituted 4-hydroxyphenyl propionic acid esters<br />
H.THOMAS ET AL. 319<br />
hepatomegaly was accompanied by an induction <strong>of</strong> total microsomal<br />
cytochrome P450, microsomal epoxide hydrolase and UDPglucuronosyltransferase,<br />
cytosolic glutathione S-transferase and<br />
peroxisomal β-oxidation. The observed induction <strong>of</strong> the cytochrome P450<br />
content was reflected in increased activities <strong>of</strong> ethoxycoumarin O-deethylase<br />
as well as lauric acid 11- and 12-hydroxylases (Table 23.3). This<br />
suggests an induction <strong>of</strong> cytochrome P450 isoenzymes <strong>of</strong> the subfamilies<br />
CYP2B and CYP4A. Indeed, increased amounts <strong>of</strong> CYP2B and CYP4A<br />
proteins were found in liver microsomes from treated animals by means <strong>of</strong><br />
Western-blotting with monoclonal antibodies specific for these iso-enzymes<br />
(data not shown). Within 28 days after cessation <strong>of</strong> a 14-day treatment at
320 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
Table 23.2 Parent equivalents and blood metabolites <strong>of</strong> [ 14 C]-labelled Compound B<br />
after single oral administration <strong>of</strong> 9.5 mg kg −1 body weight to male rats<br />
Note:<br />
bld: below the limit <strong>of</strong> detection.<br />
Blood was taken at the indicated time intervals and extracted with ethyl acetate for<br />
analysis. Compound B and its free carboxylic acid metabolite (Compound A,<br />
Table 23.1) were identified by thin-layer co-chromatography with authentic<br />
reference samples. Quantification <strong>of</strong> <strong>Compounds</strong> A and B was accomplished by<br />
radiometric scanning <strong>of</strong> the plates following thin-layer chromatography <strong>of</strong> the<br />
respective blood extracts.<br />
1000 ppm, liver weights as well as the investigated biochemical parameters<br />
returned to control levels. Therefore, Compound B may be addressed as a<br />
reversible barbiturate- and peroxisome proliferator-type inducer in the rat,<br />
as characterised by its liver enzyme induction pr<strong>of</strong>ile.<br />
Effects on serum thyrotropin and thyroid hormones<br />
When male rats were fed Compound B admixed in the diet for 14 days at<br />
dose levels <strong>of</strong> 50, 150, 500 and 1000 ppm, liver and thyroid weights were<br />
increased in a dose-dependent manner. Histopathological examination <strong>of</strong><br />
the thyroid gland revealed hypertrophy <strong>of</strong> the follicular epithelium and<br />
thinning <strong>of</strong> colloid at 150, 500 and 1000 ppm. Morphological changes in<br />
the pituitary gland comprised enlarged thyrotropin (TSH)-producing cells<br />
with foamy or vacuolated cytoplasm. In addition, treatment resulted in<br />
markedly increased serum TSH and reverse triiodothyronine (rT 3)<br />
concentrations, whereas serum thyroxine (T 4) and triiodothyronine (T 3)<br />
levels were found slightly decreased. The effects observed at 1000 ppm<br />
were found to be reversible after a 28-day recovery period (Muakkassah-<br />
Kelly et al., 1991).<br />
In additional experiments, male rats were rendered hypothyroid, fed<br />
Compound B at 1000 ppm for 21 days and infused for the last 7 days with<br />
slightly supraphysiological concentrations <strong>of</strong> T 4. The observed changes in
H.THOMAS ET AL. 321<br />
Table 23.3 The effect <strong>of</strong> Compound B on selected biochemical parameters in the<br />
male rat liver<br />
Notes:<br />
Values are means <strong>of</strong> 9 (14 days treatment) or 4 (14 days treatment/28-days<br />
recovery) animals.<br />
Standard deviations are given in parentheses.<br />
0/0: Control recovery. 1000/0: High-dose recovery.<br />
* p
322 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
metabolism has been observed with various deiodinase inhibitors (Hill et<br />
al., Liang et al., 1993) and hepatic enzyme inducers (McClain, 1989;<br />
Curran and DeGroot, 1991; Barter and Klaassen, 1992; Visser et al.,<br />
1993).<br />
Induction <strong>of</strong> thyroid neoplasia<br />
In a long-term feeding study, administration <strong>of</strong> Compound B to rats at a<br />
dose level <strong>of</strong> 1000 ppm was associated with an increased incidence <strong>of</strong><br />
thyroid gland follicular adenoma and carcinoma (Muakkassah-Kelly et al.,<br />
1991). Compound B was shown to be devoid <strong>of</strong> mutagenic and clastogenic<br />
activity (Muakkassah-Kelly et al., 1991). Therefore, it is likely that thyroid<br />
tumour induction by Compound B was not the result <strong>of</strong> a direct, genotoxic<br />
effect on this organ, but rather a consequence <strong>of</strong> the hormonal imbalance<br />
induced by this antioxidant in the rat.<br />
An intact hypothalamic-pituitary-thyroid axis is able to respond to a<br />
chemically induced alteration in peripheral hormone metabolism with<br />
increased hormone production by the hypertrophic gland. However, it is<br />
known that chronic, excessive stimulation <strong>of</strong> the gland can lead to<br />
follicular hyperplasia and ultimately progress to thyroid neoplasia (Paynter<br />
et al., 1988; McClain, 1989; Curran and DeGroot, 1991; Johnson et al.,<br />
1993). For Compound B, this hypothesis is in agreement with the doseresponse<br />
characteristics obtained in the long-term study, where thyroid<br />
tumours were induced exclusively at a dose-level sufficiently high to cause<br />
hormonal imbalance (Muakkassah-Kelly et al., 1991).<br />
Implications for human risk assessment<br />
For human risk assessment, it is <strong>of</strong> critical importance to identify the<br />
mechanism by which Compound B caused thyroid neoplasia in the rat, a<br />
species most <strong>of</strong>ten used for carcinogenic hazard identification. Hyperplastic<br />
changes in the thyroid are frequently observed in rat carcinogenicity<br />
studies, and this species appears to be very sensitive to compounds which<br />
interfere with thyroid hormone synthesis and/or catabolism. They evoke an<br />
immediate stimulation <strong>of</strong> the gland upon short-term treatment as a<br />
consequence <strong>of</strong> an increased pituitary TSH secretion (Zbinden, 1987;<br />
Paynter et al., 1988; McClain, 1989).<br />
However, the effects <strong>of</strong> xenobiotics on the pituitary-thyroid axis in<br />
rodents cannot necessarily be extrapolated to man since rodents and man<br />
are distinguished by many important physiological and biochemical<br />
differences (Gopinath, 1991): e.g. the amount <strong>of</strong> thyroxin-binding<br />
globulin, the half-life <strong>of</strong> T 4 and its biliary excretion as well as plasma THSlevels<br />
and their response to thyrotropin releasing hormone. These<br />
differences render the rat very sensitive to small changes in the plasma T 4
level, whilst humans are essentially insensitive (Hill et al., 1989; Grasso et<br />
al., 1991).<br />
Therefore, in contrast to rats, there is no conclusive evidence for a critical<br />
role <strong>of</strong> TSH in thyroid stimulation and carcinogenesis in humans (Hill et<br />
al., 1989). In cultured human thyroid cells, for example, THS was unable<br />
to induce proliferation whereas a stimulation <strong>of</strong> growth was observed in<br />
rat thyroid cells (Westermark et al., 1985). Clinical data are available for<br />
some compounds such as the anticonvulsants phenobarbitone,<br />
diphenylhydantoin and carbamazepine as well as the antibiotic rifampicin,<br />
which are known liver microsomal enzyme inducers in man. They increase<br />
thyroid hormone metabolism and excretion and eventually decrease serum<br />
thyroid hormone levels. There is also evidence, that administration <strong>of</strong> these<br />
drugs leads to thyroid stimulation, however, largely in the absence <strong>of</strong><br />
increased TSH levels (Curran and DeGroot, 1991). In addition,<br />
epidemiological data are not in favour <strong>of</strong> a link between human use <strong>of</strong> such<br />
compounds with an increased incidence <strong>of</strong> thyroid tumours (Curran and<br />
DeGroot, 1991), nor have increased rates <strong>of</strong> thyroid cancers been reported<br />
in areas <strong>of</strong> endemic iodine deficiency (McClain, 1989). Therefore, the<br />
currently available data do not support the idea, that thyroid stimulation<br />
as a response to chemically induced increases in circulating TSH<br />
concentrations significantly contributes to thyroid tumour formation in<br />
man.<br />
Compound B has been shown to cause thyroid tumours in the rat. In a<br />
series <strong>of</strong> speciality studies, the compound was identified as an enzyme<br />
inducer and a 5′-deiodinase inhibitor in the rat liver. These findings argue<br />
for a rodentspecific, indirect mechanism leading to the formation <strong>of</strong> thyroid<br />
tumours. Moreover, the observed dose-response characteristics are<br />
indicative <strong>of</strong> a threshold process, e.g. liver enzyme induction and inhibition<br />
<strong>of</strong> 5′-deiodination are irrevocable prerequisities for thyroid tumour<br />
formation in this species.<br />
Benzotriazole-based light stabilisers<br />
H.THOMAS ET AL. 323<br />
Ester derivatives <strong>of</strong> the 3-[3-(2H-benzotriazole-2-yl)-5-tert-butyl-4hydroxy-phenyl]<br />
propionic acid represent potent UV-light absorbers and<br />
constitute an important class <strong>of</strong> industrial plastic additives and light<br />
stabilisers (<strong>Compounds</strong> C-F, Table 23.1). Toxicologically, this class <strong>of</strong><br />
chemicals is characterised by generally low acute oral or dermal toxicity<br />
and the lack <strong>of</strong> genotoxicity in the commonly employed battery <strong>of</strong><br />
bacterial and cellular mutagenicity tests. Irrespective <strong>of</strong> the alcohol moiety,<br />
however, all compounds, when administered subchronically to rats,<br />
displayed very similar predominantly hepatotrophic effects, with spleen and<br />
kidney weights in addition being only slightly affected: pronounced<br />
hepatomegaly, hepatocyte hypertrophy, and concomitantly increased
324 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
plasma transaminase activities. Upon electron microscopical examination,<br />
a striking peroxisome proliferation was the major finding.<br />
Compound F, a di-ester ‘product by process’ obtained upon esterification<br />
<strong>of</strong> 3-[3-(2H-benzotriazole-2-yl)-5-tert-buty1–4-hydroxyphenyl] propionic<br />
acid with polyethyleneglycol 300, when tested for its toxicity to rat<br />
reproduction in a Segment I study, gave rise to increased numbers <strong>of</strong> stillborn<br />
pups, decreased pup survival, decreased weight gain <strong>of</strong> surviving pups, and<br />
dark discoloured abdominal skin regions in a number <strong>of</strong> pups at higher<br />
dose levels.<br />
The common nature <strong>of</strong> general toxicology findings with all investigated<br />
derivatives <strong>of</strong> the addressed benzotriazole-based light stabilisers suggested a<br />
common basis <strong>of</strong> action and, depending upon this action, perhaps a very<br />
similar behaviour and extent <strong>of</strong> potency as foetotoxic agents. In order to<br />
investigate these interrelationships a series <strong>of</strong> mechanistic studies were<br />
conducted focusing on the kinetics, primary metabolism <strong>of</strong> the parent<br />
compounds in vitro and in vivo and their effect on selected biochemical<br />
liver parameters in rats. Compound F was selected as a model compound<br />
to investigate the mechanism <strong>of</strong> toxicity in pregnant female rats and<br />
foetuses.<br />
In vitro hydrolysis<br />
Compound D, the methyl ester <strong>of</strong> 3-[3-(2H-benzotriazole-2-yl)-5-tertbutyl-4-hydroxyphenyl]<br />
propionic acid was readily hydrolysed in vitro by<br />
rat serum as well as rat liver homogenate while a homogenate <strong>of</strong> rat small<br />
intestine when compared on a gram tissue basis, appeared to be less<br />
efficient by three orders <strong>of</strong> magnitude than the liver. Increasing the sterical<br />
hindrance around the ester bond by formation <strong>of</strong> the di-ester with a short<br />
chain alcohol reduced the rate <strong>of</strong> in vitro hydrolysis considerably as<br />
demonstrated for Compound F, the diester <strong>of</strong> hexane-l,6-diol with<br />
Compound C. When <strong>of</strong>fered at a test concentration <strong>of</strong> 0.2 mM, essentially<br />
no hydrolysis <strong>of</strong> this compound was observed with rat serum, and the<br />
hydrolysis by liver and small intestine homogenate was estimated to<br />
proceed at least two and one orders <strong>of</strong> magnitude slower, respectively, than<br />
calculated for Compound D (Table 23.1 and 23.4).<br />
Blood kinetics and blood metabolites<br />
Assuming comparable extents <strong>of</strong> intestinal absorption in vivo, the observed<br />
differences in the in vitro hydrolysis rates might as well suggest<br />
significantly different in vivo hydrolysis rates and consequently quite<br />
different residence times for both parent compounds in the rat in vivo.<br />
Different residence times, on the other hand, may eventually allow not only<br />
for additional routes <strong>of</strong> metabolism but also for an intensification <strong>of</strong> toxic
Table 23.4 Kinetic parameters for the in vitro hydrolysis <strong>of</strong> Compound D and E by<br />
rat serum and organ homogenates<br />
Notes:<br />
a The apparent Vmax value for serum is given in µmol min −1 ml −1.<br />
H.THOMAS ET AL. 325<br />
b Initial velocity <strong>of</strong> hydrolysis at 0.2 mM ester concentration.<br />
Hydrolysis <strong>of</strong> Compound D was determined in 50 mM Tris/phosphate buffer, pH 7.<br />
5, containing either 1 per cent (v/v) rat serum, or 1.25 per cent (w/v) rat liver<br />
homogenate or 10 per cent (w/v) small intestine homogenate. Similarly, hydrolysis<br />
<strong>of</strong> Compound E was assessed using 98 per cent (v/v) rat serum or in the presence <strong>of</strong><br />
10 mM Tris/HCl buffer, pH 7.5, containing 250 mM sucrose and either 24.5 per<br />
cent (w/v) rat liver homogenate or 19.6 per cent (w/v) small intestine homogenate.<br />
effects. This question was addressed in a pharmacokinetic study under<br />
conditions <strong>of</strong> single oral administration <strong>of</strong> <strong>Compounds</strong> D and E at a dose<br />
level <strong>of</strong> 10 mg kg −1 each (Table 23.5).<br />
14 C-Phenyl-labelled Compound D was readily absorbed from the<br />
gastrointestinal tract. Maximal blood radioactivity was reached between 1<br />
and 2 h and subsequently eliminated with an apparent half-life <strong>of</strong> 10.0–11.<br />
8 h. After 48 h only minute amounts <strong>of</strong> radioactivity equalling about 3 per<br />
cent <strong>of</strong> the blood levels in T max were detectable. Analysis <strong>of</strong> the resulting<br />
blood metabolite pattern largely confirmed the findings <strong>of</strong> the preceding in<br />
vitro investigation <strong>of</strong> enzymatic Compound D hydrolysis: particularly<br />
during periods <strong>of</strong> high parent equivalent concentrations in blood as<br />
recorded between 30 min and 6 h after dosing, hydrolysis appeared to be<br />
the major metabolic pathway as evidenced by the high concentrations <strong>of</strong><br />
the carboxylic acid, Compound C (34–77 per cent), and an unidentified<br />
metabolite (17–36 per cent) which was regarded to have evolved from<br />
Compound C by an additional metabolic step (Table 23.5).<br />
Quite surprisingly, Compound E was found to be absorbed to a much<br />
lower extent than Compound D, with a C max after 1 h <strong>of</strong> less than one<br />
tenth <strong>of</strong> the value seen with the latter. Elimination with an apparent halflife<br />
<strong>of</strong> 12.0 h and slightly higher residual radioactivity <strong>of</strong> approximately 4.8<br />
per cent <strong>of</strong> the blood levels recorded at T max after 48 h indicated only a<br />
slightly reduced elimination rate as compared to Compound D. Also, quite<br />
different from the initially anticipated result, hydrolysis contributed<br />
substantially to the rapid metabolism <strong>of</strong> Compound E. The 24 h AUC<br />
values revealed a 23 per cent and 36 per cent contribution <strong>of</strong> the carboxylic
326 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
Table 23.5 Summary <strong>of</strong> pharmacokinetic parameters obtained after single oral<br />
administration <strong>of</strong> 10 mg kg −1 Compound D and E to two male rats each<br />
Notes:<br />
a Below the limit <strong>of</strong> reliable quantification.<br />
pe: Parent equivalent.<br />
acid, Compound C, and unidentified metabolites, respectively, to the total<br />
AUC. It is assumed that the majority <strong>of</strong> unidentified metabolites have<br />
arisen from further biotransformation <strong>of</strong> the carboxylic acid. The<br />
significant contribution <strong>of</strong> hydrolysis products to the total AUC is still<br />
evident after 168 h although low blood radioactivity levels 24 h after<br />
dosing generally prevented accurate quantitation <strong>of</strong> the metabolites and<br />
thus slightly diminished their apparent overall share (Table 23.5).<br />
Liver enzyme induction<br />
Subchronic oral (gavage) administration <strong>of</strong> single daily doses <strong>of</strong> Compound<br />
C for 14 days, Compound D for 14 days, Compound E for 13 weeks, and<br />
Compound F for 114 days to male rats (Tables 23.6 and 23.7) was<br />
correlated with a dose-dependent massive increase in absolute liver weight<br />
up to about 190 per cent <strong>of</strong> control at the highest dose level irrespective <strong>of</strong><br />
the treatment period and the test compound. This pronounced<br />
hepatomegaly was paralleled by a comparably small two-fold elevation <strong>of</strong><br />
the microsomal cytochrome P450 contents and an about 50 per cent<br />
decrease in total UDP- glucuronosyltransferase activity. Essentially no<br />
changes were recorded for the cytochrome P450 dependent<br />
ethoxycoumarin O-de-ethylase activity while microsomal epoxide<br />
hydrolase activities appeared to vary, with slight increases in Compound C<br />
and Compound D treated animals, no changes in Compound E treated<br />
animals and even a dose-dependent reduction to 46 per cent <strong>of</strong> control in<br />
rats treated with Compound F at 100 mg kg −1 (Table 23.6).<br />
Strongly induced peroxisomal fatty acid β-oxidation activities for all<br />
model compounds as well as lauric acid 12-hydroxylase activities tested for<br />
<strong>Compounds</strong> E and F were accompanied by significant dose-dependent<br />
decreases in glutathione S-transferase activities to 32, 51, 40 and 15 per<br />
cent <strong>of</strong> control at the highest dose level tested for Compound C, D, E and
Table 23.6 The effect <strong>of</strong> various benzotriazole-based light stabilisers on selected biochemical parameters related to and indicative<br />
for a barbiturate and/or polycyclic aromatic hydrocarbon type enzyme induction in male rat liver<br />
H.THOMAS ET AL.<br />
327<br />
Notes:<br />
dnt: dose level not tested,<br />
nd: not determined.<br />
Microsomal epoxide hydrolase and total microsomal UDP-glucuronosyltransferase activities were determined with styrene oxide<br />
and 3-methyl-2-nitrophenol as substrate, respectively.<br />
Values are means±standard deviation <strong>of</strong> 10 control and 5 treated animals (a, b) or 8 animals (c) or 5 animals per group (d).<br />
Asterisks indicate results significantly different (two-sided Dunnett’s test) from control: * p
328 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
F, respectively. For Compound F, which appeared to be the most potent<br />
inducer <strong>of</strong> peroxisomal β-oxidation and lauric acid 12-hydroxylase<br />
activities, a concomitant strong 90 per cent reduction <strong>of</strong> morphine UDPglucuronosyltrasferase<br />
and marked 2.5-fold increase in bilirubin UDPglucuronosyltransferase<br />
activity was recorded (Table 23.7). Electron<br />
microscopy confirmed what had already been indicated by changes in the<br />
investigated enzyme levels, a striking proliferation <strong>of</strong> peroxisomes with the<br />
same appearance <strong>of</strong> these organelles regardless <strong>of</strong> the compound tested:<br />
vigorous increase in number, a striking number <strong>of</strong> markedly enlarged<br />
peroxisomes frequently containing matrical inclusions (matrical plates) and<br />
peroxisomes forming arrays or clusters (polyperoxisomes) in the virtual<br />
absence <strong>of</strong> any significant proliferation <strong>of</strong> smooth endoplasmic reticulum<br />
(data not shown).<br />
Consequently, the hepatotrophic effects <strong>of</strong> the tested benzotriazole-based<br />
light stabilisers were clearly assigned to their action as peroxisome<br />
proliferators in rat liver. Mindful <strong>of</strong> the different durations <strong>of</strong> treatment<br />
their potency was found to rank in the order Compound E
Table 23.7 The effect <strong>of</strong> various benzotriazole-based light stabilizers on selected biochemical parameters related to and indicative<br />
for a peroxisome proliferator type enzyme induction in male rat liver<br />
H.THOMAS ET AL. 329<br />
Notes:<br />
dnt: dose level not tested.<br />
nd: not determined.<br />
Cyanide-insensitive peroxisomal fatty acid -oxidation and cytosolic glutathione S-transferase activities were determined with<br />
[l-14C]-palmitoyl-CoA and1-chloro2,4-dinitrobenzene as substrate, respectively.<br />
Values are means±standard deviation <strong>of</strong> 10 control and 5 treated animals (a, b) or 8 animals (c) or 5 animals per group (d).<br />
Asterisks indicate results significantly different (two-sided Dunnett’s test) from control: * p
330<br />
ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
Table 23.8 Morphological changes <strong>of</strong> dam and foetal hepatocyte organelles after treatment with Compound F from day 6 through<br />
days 14, 17 and 20 <strong>of</strong> gestation
iochemical investigations which revealed a moderate four-fold induction<br />
<strong>of</strong> peroxisomal β-oxidation in dams and an up to 15-fold induction <strong>of</strong> this<br />
activity at day 21 <strong>of</strong> gestation in foetal livers, respectively, with the final<br />
foetal activity exceeding dam activity by 40 per cent. Strong increases in<br />
lauric acid 11- and 12-hydroxylation, which are known to be associated<br />
with isoenzymes <strong>of</strong> the cytochrome P450 CYP4A gene family and the<br />
phenomenon <strong>of</strong> peroxisome proliferation in rodents, <strong>of</strong> up to 2.6- and 10.<br />
5-fold in dams and 11.5- and 23.2-fold in foetuses, respectively, at day 21<br />
<strong>of</strong> gestation were recorded as well, indicating again a slightly higher final<br />
activity in foetal than in dam liver. By contrast, catalase, known as a<br />
detoxifying enzyme for hydrogen peroxide generated particularly in the<br />
course <strong>of</strong> increased peroxisomal activity, was shown to be induced up to 5.<br />
6-fold in dam and 8.2-fold in foetal liver at day 21 <strong>of</strong> gestation, leaving<br />
foetal liver, however, with only about 30 per cent <strong>of</strong> the final activity seen<br />
in dams (Table 23.9). This discrepancy between the strong induction <strong>of</strong><br />
hydrogen peroxide generating peroxisomal β-oxidation and the<br />
considerably less potent induction <strong>of</strong> the hydrogen peroxide destroying<br />
catalase activity appears to be reflected in the 1.7-fold increase in the lipid<br />
peroxidation product malondialdehyde in foetal livers as well as in the<br />
decrease <strong>of</strong> total and reduced hepatic glutathione levels at day 21 <strong>of</strong><br />
gestation (Tables 23.9 and 23.10). Surprisingly, foetal defense systems<br />
against oxidative stress such as selenium-dependent and<br />
seleniumindependent glutathione peroxidase were found poorly developed<br />
throughout the investigated periods <strong>of</strong> gestation and barely inducible by<br />
the test article leaving the pups with little protection against any kind <strong>of</strong><br />
oxidative insult (Table 23.10). Consequently, Compound F was clearly<br />
identified as a peroxisome proliferator in pregnant rat as well as foetal liver<br />
with high potential for the initiation <strong>of</strong> oxidative damage in foetal tissues.<br />
Implications for human safety assessment<br />
H.THOMAS ET AL. 331<br />
The understanding <strong>of</strong> the mechanisms by which benzotriazole-based UV<br />
light stabilisers exert their hepatotrophic effects in rodents is <strong>of</strong> crucial<br />
importance for the assessment <strong>of</strong> safety aspects in humans. The presented<br />
rat studies have shown that the liver effects exerted by Compound C and<br />
its ester derivatives <strong>Compounds</strong> D, E and F are clearly related to the<br />
induction <strong>of</strong> peroxisome proliferation. The identical nature <strong>of</strong> the observed<br />
effects regardless <strong>of</strong> the alcohol component in the investigated esters<br />
suggests that the toxic potential resides solely with the 3-[3-(2Hbenzotriazole-2-yl)-5-tert-butyl-4-hydroxy-phenyl]<br />
propionic acid whereby<br />
the individual potency appears to be mainly determined by the different<br />
bioavailablity <strong>of</strong> the respective ester compound and the extent and velocity<br />
<strong>of</strong> its hydrolysis in vivo. Also, it appears, that in vitro hydrolysis studies do
332 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
not actually reflect the hydrolytic capacity <strong>of</strong> the in vivo system for a given<br />
ester.<br />
Liver enlargement and the induction <strong>of</strong> diagnostic enzyme activities is a<br />
characteristic response <strong>of</strong> rodents to treatment with peroxisome<br />
proliferators and results from a combination <strong>of</strong> both hypertrophy and<br />
hyperplasia. According to current opinion, peroxisome proliferation and<br />
liver growth are closely associated with the formation <strong>of</strong> hepatocellular<br />
tumours in rats and mice (Hawkins et al., 1987; Lock et al., 1989; Bentley<br />
et al., 1993). However, a number <strong>of</strong> feeding studies have demonstrated<br />
that there may actually be two types <strong>of</strong> threshold with respect to dose<br />
relationships: at very low doses, administration <strong>of</strong> peroxisome proliferators<br />
will not result in any liver response at all (Bentley et al., 1993). With<br />
increasing doses the first threshold will be exceeded with subsequent<br />
stimulation <strong>of</strong> peroxisome proliferation and DNA synthesis.<br />
As a result <strong>of</strong> several studies it is obvious that a limited extent <strong>of</strong> liver<br />
growth does not automatically lead to tumour formation as has been<br />
demonstrated, for example, with fen<strong>of</strong>ibrate and diethylhexylphthalate in<br />
carcinogenicity assays (Mitchell et al., 1985; Price et al., 1986; Keith et al.,<br />
1991). Thus, a second threshold has to be exceeded at which the<br />
magnitude <strong>of</strong> effects is sufficient to cause tumour development in rodents.<br />
In addition, extended administration <strong>of</strong> the peroxisome proliferator<br />
appears a necessary prerequisite to exceed this tumourigenic threshold.<br />
Also, a large number <strong>of</strong> in vitro and in vivo studies have provided ample<br />
evidence for a marked species difference in susceptibility to the effects <strong>of</strong><br />
peroxisome proliferators. Rats and mice are extremely sensitive while<br />
hamsters show a markedly smaller response and non-human primates and<br />
humans appear to be insensitive or non-responsive (Lake et al., 1989;<br />
Bentley et al., 1993; Graham et al., 1994). The latter finding is supported<br />
by epidemiological evidence from long-term treatment <strong>of</strong> patients with<br />
hypolipidaemic agents (Bentley et al., 1993). Therefore the available<br />
evidence strongly supports the conclusion that the effects <strong>of</strong> benzotriazolebased<br />
light stabilisers in rodents are <strong>of</strong> no relevance to human safety<br />
assessment.<br />
The action <strong>of</strong> Compound F as a strong peroxisome proliferator in foetal<br />
livers starting as early as day 15 <strong>of</strong> gestation suggests that treatment related<br />
initiation <strong>of</strong> high level oxidative stress under conditions <strong>of</strong> poorly<br />
developed foetal protection and detoxification systems. This view is<br />
supported by substantially elevated hepatic malondialdehyde levels and<br />
essentially depleted glycogen stores in hepatocytes <strong>of</strong> foetuses from treated<br />
dams on day 21 <strong>of</strong> gestation. The latter indicates an extensive glucose<br />
consumption, presumably via the pentose phosphate pathway, to supply<br />
the hydrogen peroxide and lipid peroxide detoxifying glutathione<br />
peroxidase system with the necessary reduction equivalents (NADPH).<br />
Under conditions <strong>of</strong> limited degradation <strong>of</strong> hydrogen peroxide, this
Table 23.9 The effect <strong>of</strong> Compound F on dam and foetal absolute liver weight and selected biochemical liver parameters related to<br />
and indicative for peroxisome proliferation after treatment from day 6 through days 14, 17 and 20 <strong>of</strong> gestation<br />
H.THOMAS ET AL. 333<br />
Notes:<br />
bld: below the limit <strong>of</strong> detection.<br />
Cyanide-insensitive peroxisomal fatty acid -oxidation and catalase activities were determined with [lhydrogen<br />
peroxide as substrate, respectively.<br />
C]-palmitoyl-CoA and<br />
Values are means±standard deviation from 6 dams per group and 6 pools <strong>of</strong> 2–9 foetuses per dam depending on the litter size.<br />
Asterisks indicate results significantly different (two-sided Dunnett’s test) from control: * p
334<br />
ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
Table 23.10 The effect <strong>of</strong> Compound F on selected biochemical dam and foetal liver parameters related to defence mechanisms<br />
against oxidative stress after treatment from day 6 through days 14, 17 and 20 <strong>of</strong> gestation<br />
Notes:<br />
Selenium-dependent and selenium-independent glutathione peroxidase activities were determined with cumene hydroperoxide and<br />
hydrogen per-oxide as substrate, respectively.<br />
Values are means±standard deviation from 6 dams per group and 6 pools <strong>of</strong> 2–9 foetuses per dam depending on the litter size.<br />
Asterisks indicate results significantly different (two-sided Dunnett’s test) from control: * p
compound is known to pass the liver and to be systemically distributed<br />
throughout the foetal body. Thus the endothelial cell injuries observed in<br />
the course <strong>of</strong> the Segment II study are regarded secondary to the effect <strong>of</strong><br />
peroxisome proliferation and to have arisen as a consequence <strong>of</strong> oxidative<br />
damage which has been demonstrated to occur in various endothelial cell<br />
systems, for example as a result <strong>of</strong> iron-mediated oxygen free radical attack<br />
(Brieland et al., 1992; Barchowsky et al., 1994; Krautschick et al., 1995).<br />
Therefore, the observed foetotoxicity <strong>of</strong> Compound F in rats is regarded as<br />
occurring solely as a consequence <strong>of</strong> and secondary to peroxisome<br />
proliferation, and there is no evidence to assume that this effect as a result<br />
<strong>of</strong> exposure to benzotriazole based light stabilisers may occur in species<br />
that are non-responsive to the action <strong>of</strong> peroxisome proliferators as stated<br />
above, including humans.<br />
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J.C. S., 1989, Butylated hydroxyanisole-induced alterations in cell kinetic<br />
parameters in rat forestomach in relation to its oxidative cytochrome P-450mediated<br />
metabolism, Carcinogenesis, 10, 1947–51.<br />
VERHAGEN, H., SCHILDERMAN, P.A.E.L. and KLEINJANS, J.C.S., 1991,<br />
Butylated hydroxyanisole in perspective, Chemico-Biological Interactions, 80,<br />
109– 34.<br />
VISSER, T.J., KAPTEIN, E., VAN TOOR H., VAN RAAIJ, J.A.G.M., VAN DEN<br />
BERG, K.J., TJONG TJIN JOE, C., VAN ENGELEN, J.G.M. and BROUWER<br />
A., 1993, Glucuronidation <strong>of</strong> thyroid hormone in rat liver: effects <strong>of</strong> in vivo
338 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT<br />
treatment with microsomal enzyme inducers and in vitro assay conditions,<br />
Endocrinology, 133, 2177–86.<br />
WESTERMARK, K., KARLSSON, F.A. and WESTERMARK, B., 1985,<br />
Thyrotropin modulates EGF receptor function in porcine thyroid follicle cells,<br />
Molecular and Cellular Endocrinology, 40, 17–23.<br />
WILLIAMS, G.M., McQUEEN, C.A. and TONG, C., 1990a, Toxicity studies <strong>of</strong><br />
butylated hydroxyanisole and butylated hydroxytoluene. I. Genetic and<br />
cellular effects, Food and Chemical <strong>Toxicology</strong>, 28, 793–8.<br />
WILLIAMS, G.M., WANG, C.X. and IATROPOULOS, M.J., 1990b, Toxicity<br />
studies <strong>of</strong> butylated hydroxyanisole and butylated hydroxytoluene, II. Chronic<br />
feeding studies. Food and Chemical <strong>Toxicology</strong>, 28, 799–806.<br />
ZBINDEN, G., 1987, Assessment <strong>of</strong> hyperplastic and neoplastic lesions <strong>of</strong> the<br />
thyroid gland, Trends in Pharmacological Sciences, 8, 11–14.
24<br />
<strong>Toxicology</strong> <strong>of</strong> Surfactants: Molecular,<br />
Mechanistic and Regulatory Aspects<br />
WALTER STERZEL<br />
Henkel KGaA, Düsseldorf<br />
Introduction<br />
The vast distribution <strong>of</strong> surfactants in various products in everyday use<br />
requires that the unwanted effects as well as the desired properties are<br />
known in order to recognize possible risks and prevent any damage to the<br />
health <strong>of</strong> humans. In order to understand the effects <strong>of</strong> surfactants on the<br />
organism, their most important biochemical effects, which depend on the<br />
interaction <strong>of</strong> surface active agents with basic biological structures like<br />
membranes, proteins and enzymes, are discussed. Following this discussion<br />
the local effects <strong>of</strong> surfactants are described. Local in this sense are all the<br />
effects encountered directly at the point <strong>of</strong> contact with the outer surfaces<br />
<strong>of</strong> the body, such as skin and mucous membrane irritation as well as<br />
allergies arising from skin contact. After this section the toxicokinetic<br />
properties <strong>of</strong> surfactants, providing information about type and extent <strong>of</strong><br />
absorption by organisms, metabolic pathways and their elimination are<br />
discussed. The section on systemic effects deals, in contrast to local effects,<br />
with reactions arising after the substance has entered the organism by<br />
swallowing, skin penetration or inhalation.<br />
Due to their technical and economic importance, surfactants have been<br />
used extensively for decades. This resulted in an abundance <strong>of</strong> scientific<br />
publications concerning their effects on organisms. As a complete review<br />
would exceed the scope <strong>of</strong> this contribution, the focus will centre on a<br />
description <strong>of</strong> exemplary data which are important for the evaluation <strong>of</strong><br />
the safety <strong>of</strong> surfactants.<br />
Biochemical properties <strong>of</strong> surfactants<br />
Surfactants come into immediate contact with the body during cleaning <strong>of</strong><br />
the skin and act on the skin cells directly. When surfactants are swallowed<br />
unin tentionally, tissue damage is also possible. The question <strong>of</strong> the effects<br />
on the cells and cell components like membranes, proteins and enzymes is<br />
therefore also important from a toxicological point <strong>of</strong> view.
340 TOXICOLOGY OF SURFACTANTS<br />
Interactions with membranes<br />
Due to their ability to absorb at interfaces, surfactants can interact with<br />
biological membranes. This interaction depends on the concentration <strong>of</strong> the<br />
surfactant and can be described in the following sequence (Helenius and<br />
Simons, 1975). In the first instance the monomeric surfactant molecule<br />
adsorbs onto the membrane. For a low surfactant/membrane ratio this<br />
changes the permeability <strong>of</strong> the membrane and leads to cell lysis at higher<br />
concentrations. At even higher surfactant concentrations, the lamellar<br />
structure <strong>of</strong> the membrane is lost and it is solubilized. A further increase in<br />
surfactant concentration results in the separation <strong>of</strong> the phospholipids from<br />
the protein. This allows surfactant molecules to adsorb on previously<br />
hidden regions <strong>of</strong> the protein molecule. For the solubilization <strong>of</strong> integral<br />
membrane proteins the formation <strong>of</strong> micelle/protein complexes seems to be<br />
a prerequisite. A significant solubilization <strong>of</strong> these proteins is possible only<br />
if the critical micelle concentration c M is exceeded. This is indicated by the<br />
fact that the microsomal membrane bound enzyme arylsulphatase-C could<br />
only be extracted from the membrane with retention <strong>of</strong> the biological<br />
activity after micelles were formed (Chang et al., 1985).<br />
As a consequence <strong>of</strong> these interactions, surfactants are able to influence<br />
the metabolism <strong>of</strong> membrane components (DeLeo, 1989). This has been<br />
demonstrated by studies on the pathophysiology <strong>of</strong> surfactant-mediated<br />
skin irritation. In vitro cultured corneocytes showed an increased release <strong>of</strong><br />
cholin metabolites after incubation with anionic surfactants. This effect<br />
was less pronounced after treatment with nonionic surfactants. In<br />
conclusion, these investigations demonstrated that the release <strong>of</strong><br />
metabolites is correlated with the irritation potential <strong>of</strong> surfactants.<br />
Interactions with proteins<br />
Depending on the structure <strong>of</strong> the surfactant the interactions with proteins<br />
are based on polar or hydrophobic interactions. The binding <strong>of</strong> surfactants<br />
to protein molecules is a function <strong>of</strong> the concentration <strong>of</strong> free surfactant in<br />
equilibrium with the protein. The binding is affected by the pH,<br />
temperature and ionic strength <strong>of</strong> the solution. These factors can lead to<br />
conformational changes <strong>of</strong> proteins and thereby increase or decrease the<br />
number <strong>of</strong> available binding sites. Natural bovine albumin, for example,<br />
has 10 binding sites for decyl glucoside at 10°C and 13 at 25°C<br />
(Wasylewski and Kozik, 1979). According to a theory developed by Jones<br />
(1975), surfactants adsorb onto proteins in multiple equilibria. Only a few<br />
surfactant molecules (
W.STERZEL 341<br />
pyruvate oxidase can form this type <strong>of</strong> binding (Schwuger and Bartnik,<br />
1980). Binding <strong>of</strong> more surfactant molecules leads to conformational<br />
changes in the protein. It is obvious that conformational changes allow<br />
binding <strong>of</strong> further surfactant molecules on hydrophobic regions which were<br />
previously not exposed.<br />
According to their different chemical structure (e.g. anionic, cationic,<br />
amphoteric or nonionic) surfactants differ significantly in their ability to<br />
carry out cooperative binding and therefore they differ in their biological<br />
activity. Anionic surfactants form adsorption complexes with proteins due<br />
to polar and hydrophobic interactions. Polar interactions between the<br />
negatively charged hydrophilic group <strong>of</strong> the surfactant and the positively<br />
charged groups <strong>of</strong> the protein molecule are the precondition for the<br />
formation <strong>of</strong> hydrophobic associations between surfactant molecule and<br />
protein molecule (Garcia-Dominguez, 1977; Schwuger and Bartnik, 1980).<br />
In the case <strong>of</strong> dodecylsulphate and tetradecylsulphate the binding results in<br />
denaturation <strong>of</strong> the proteins (Makino et al., 1973). Cationic surfactants can<br />
interact by polar and hydrophobic binding as well. Polar interactions result<br />
in electrostatic bonds between the negatively charged groups <strong>of</strong> the protein<br />
molecule and the positively charged surfactant molecule. For example, the<br />
enzyme, glucose oxidase, is deactivated by hexadecyl trimethyl ammonium<br />
bromide through formation <strong>of</strong> an ion pair between the cationic surfactant<br />
and the anionic amino acid side chain <strong>of</strong> the enzyme molecule (Tsuge,<br />
1984). Nonionic or amphoteric surfactants and proteins show either no<br />
interaction at all or interactions that are extremely weak and normally<br />
close to the limits <strong>of</strong> sensitivity <strong>of</strong> the analytical methods used. For this<br />
reason, nonionic surfactants will not dissolve sparingly soluble proteins,<br />
denature proteins, or contribute to a swelling <strong>of</strong> the epidermis. Figure 24.1<br />
shows the solubility <strong>of</strong> the protein zein, which is almost insoluble in water,<br />
and is more or less solubilized by sodium dodecyl sulphate and alkyl<br />
ethyleneglycol ether sulphates, while the nonionic ethoxylated nonylphenol<br />
is ineffective (Schwuger and Bartnik, 1980). A further reason for the poor<br />
interactions between nonionic surfactants and proteins could be that the<br />
concentration necessary for cooperative binding with the protein is not<br />
attained with nonionic surfactants due to their low critical micelle<br />
concentration c M (Makino et al., 1973).<br />
An important consequence <strong>of</strong> the interactions between anionic<br />
surfactants and proteins is the swelling <strong>of</strong> the stratum corneum <strong>of</strong> the skin.<br />
Hydrophobic interactions between surfactant chains and the protein result<br />
in pendant ionic head groups and subsequently in swelling because <strong>of</strong><br />
electrostatic repulsion between them. As the substrate matrix expands and<br />
the tertiary structure is disrupted, hydration occurs which leads to swelling<br />
(Blake-Haskins, 1986).
342 TOXICOLOGY OF SURFACTANTS<br />
Figure 24.1 Zein solubility c z <strong>of</strong> saturated solutions as a function <strong>of</strong> surfactant<br />
concentration. Zein concentration, 50 g lit −1 , mixing period, 2 h, temperature, 40°C<br />
( , sodium dodecyl sulphate; , alkyl ether sulphate (2EO); , nonyl phenol<br />
ethoxylate (9EO)).<br />
Interactions with enzymes<br />
Surfactants which are capable <strong>of</strong> massive cooperative binding, such as<br />
many anionic and cationic surfactants, induce conformational changes in<br />
the protein molecule which in general lead to loss <strong>of</strong> biological activity.<br />
The following mechanisms <strong>of</strong> enzyme inactivation by surfactants have to<br />
be considered (Ne’eman et al., 1971):<br />
1. Disruption <strong>of</strong> the quaternary structure <strong>of</strong> the enzyme when the enzyme<br />
protein consists <strong>of</strong> several subunits.
2. Induction <strong>of</strong> conformational changes in the tertiary or secondary<br />
structure <strong>of</strong> the enzyme protein.<br />
3. In the case <strong>of</strong> membrane-bound enzymes, separation <strong>of</strong> the enzyme<br />
protein from essential membrane lipids.<br />
4. Binding at active sites <strong>of</strong> the enzyme.<br />
While the effect <strong>of</strong> cationic surfactants on membranes is comparable to<br />
that <strong>of</strong> anionic surfactants, many proteins are obviously more resistant<br />
towards the denaturing activity <strong>of</strong> .cationic surfactants (Nozaki et al.,<br />
1974). Binding <strong>of</strong> tetradecyl trimethyl ammonium chloride onto bovine<br />
serum albumin and other proteins is comparable to that <strong>of</strong> sodium dodecyl<br />
sulphate. However, the cooperative binding with subsequent denaturation<br />
requires a ten-fold higher concentration <strong>of</strong> cationic surfactant. The<br />
saturation <strong>of</strong> the surfactant/protein complex is prevented by the competing<br />
formation <strong>of</strong> surfactant micelles. Contrary to the irreversibly denaturing<br />
effect <strong>of</strong> sodium dodecyl sulphate, the effect <strong>of</strong> some cationic surfactants on<br />
proteins is reversible (Nakaya et al., 1971).<br />
Local effects<br />
Skin compatibility<br />
W.STERZEL 343<br />
The damaging effects <strong>of</strong> surfactants on skin manifest themselves in dryness,<br />
roughness and scaling. In addition, symptoms <strong>of</strong> inflammation (reddening,<br />
swelling) can develop, which can result, in severe cases, in complete<br />
destruction <strong>of</strong> the tissue. All these symptoms are a result <strong>of</strong> the described<br />
biochemical properties <strong>of</strong> surfactants. The skin is defatted by the more or<br />
less pronounced property <strong>of</strong> the surfactants to emulsify lipids and thus<br />
partially or completely removing the surface film <strong>of</strong> lipids. This leads to a<br />
disturbance <strong>of</strong> the barrier function <strong>of</strong> the skin resulting in increased<br />
permeability for chemical substances and a loss <strong>of</strong> water. Anionic<br />
surfactants can cause swelling <strong>of</strong> the skin. As a result, they facilitate the<br />
transport <strong>of</strong> substances to lower layers where inflammation reactions can be<br />
induced (Scholz, 1967). The reaction <strong>of</strong> surfactants with proteins dissolves<br />
proteins out <strong>of</strong> the skin and leads to their denaturation. These changes in<br />
the matrix material have an effect on the resistance <strong>of</strong> the skin (Götte,<br />
1967) and, along with degreasing and drying, are an additional cause <strong>of</strong> an<br />
increase in skin roughness (Imokawa, 1975).<br />
The majority <strong>of</strong> the knowledge about skin compatibility <strong>of</strong> surfactants<br />
originates from studies with experimental animals, preferably rabbits.<br />
Furthermore, it is possible to evaluate new substances directly on human<br />
skin after careful exclusion <strong>of</strong> unreasonable risks. A critical overview <strong>of</strong><br />
different test methods is given by Kästner (1980). In this context, the
344 TOXICOLOGY OF SURFACTANTS<br />
problem <strong>of</strong> labelling chemical products with respect to their toxicological<br />
properties has to be addressed. Different national or international<br />
regulations, e.g. the EG directive 67/548 within the European Community,<br />
dictate that these products are labelled ‘irritating’ or ‘corrosive’ whenever<br />
exactly defined effects are observed in appropriate tests. With consumer<br />
protection in mind, exceedingly stringent test procedures have been<br />
established. These conditions frequently result in an unfavourable<br />
classification especially for surfactants. When interpreting data from such<br />
studies, it is important to consider that unrealistic conditions <strong>of</strong> exposure<br />
were involved.<br />
Since anionic surfactants are the class with the greatest economic<br />
importance, they are the best studied. No general statement is possible with<br />
regard to a classification <strong>of</strong> the various groups <strong>of</strong> anionic surfactants in<br />
order <strong>of</strong> their skin compatibility, since within each class <strong>of</strong> substances<br />
significant differences exist in their effect on skin depending on the<br />
respective structure. Opdyke et al. (1965), for example, found a decrease in<br />
the skin irritation potential <strong>of</strong> different alkyl ether sulphates with<br />
increasing level <strong>of</strong> ethoxylation. The effect <strong>of</strong> the alkyl chain length <strong>of</strong><br />
anionic surfactants was examined in different test models for soaps, alkyl<br />
sulphates, alkyl sulphonates, alkylbenzene sulphonates as well as alphaolefin<br />
sulphonates (Kästner, 1980). As shown in Table 24.1, it could be<br />
established in all cases that compounds with a saturated side chain <strong>of</strong> 10–<br />
12 C atoms exert the largest effect, or rather, have the highest potential for<br />
damage. When the results <strong>of</strong> skin compatibility tests for the most<br />
important classes <strong>of</strong> anionic surfactants are summarized, it becomes<br />
evident that the undiluted products have to be regarded as strongly<br />
irritating substances. Even at concentrations <strong>of</strong> 10 per cent moderate to<br />
strong effects have to be expected. However, at concentrations less than 1<br />
per cent, which is the range corresponding to typical use levels in<br />
detergents, only minimal irritation is observed.<br />
Nonionic surfactants have a good skin compatibility at normal use<br />
levels. Although studies with alcohol ethoxylates were reported in which a<br />
strong irritant effect was observed (Grupp et al., 1960), these studies used<br />
concentrations far above the usual exposure levels <strong>of</strong> consumers.<br />
Independent <strong>of</strong> their structure, cationic surfactants cause severe skin<br />
damage in high concentrations, while typical application levels are<br />
generally tolerated well.<br />
Mucous membrane compatibility<br />
When talking about mucous membrane compatibility one has to consider<br />
not only the mucous membranes <strong>of</strong> the eye. In addition, the mucous<br />
membranes in the mouth, upper and lower gastrointestinal tract as well as<br />
the urogenital tract have to be considered. In general, the effects <strong>of</strong>
Table 24.1 Structure/activity relationships <strong>of</strong> anionic surfactants<br />
W.STERZEL 345<br />
a Test model: A=epicutaneous, mouse; B=intracutaneous, mouse; C=epicutaneous,<br />
man; D=roughness <strong>of</strong> skin; E=swelling <strong>of</strong> collagen in vitro; F=denaturation <strong>of</strong><br />
protein in vitro.<br />
b Number <strong>of</strong> carbon atoms in the alkyl chain.<br />
surfactants on mucous membranes are based on the same biochemical<br />
mechanisms that are described in the chapter on skin compatibility. Special<br />
characteristics in the fine structure <strong>of</strong> mucous membranes, like the absence<br />
<strong>of</strong> keratin, result in a significantly higher sensitivity <strong>of</strong> these tissues towards<br />
chemical substances. Irritating materials affecting the eye cause reddening<br />
through increased blood flow in the conjunctivae with enlargement <strong>of</strong> the<br />
blood vessels. This can finally lead to the destruction <strong>of</strong> the cell walls<br />
accompanied by bleeding. Depending on the severity <strong>of</strong> the effects, a more<br />
or less pronounced swelling or reflex-induced closure <strong>of</strong> the eyelid will<br />
occur, followed by tearing and secretion. If the degree <strong>of</strong> irritation is low,<br />
epithelium damage develops on the cornea which can be visualized only<br />
with special techniques (staining, slit lamp microscope) and which is<br />
generally reversible. In severe cases the effects result in irreversible clouding<br />
<strong>of</strong> the cornea and therefore lead to an impairment <strong>of</strong> the eyesight.<br />
The classical method for the evaluation <strong>of</strong> mucous membrane<br />
compatibility <strong>of</strong> chemicals is the so-called Draize test on the rabbit eye<br />
(Draize et al., 1944). A structure/activity relationship with respect to the<br />
length <strong>of</strong> the respective alkyl chains <strong>of</strong> anionic surfactants can, as for the<br />
skin compatibility, also be observed for the mucous membrane<br />
compatibility (Kästner, 1980). According to this, the maximum irritation<br />
occurs at chain lengths <strong>of</strong> C 10−14. for n-alkyl sulphates as well as for nalkyl<br />
sulphonates. Although the irritation potential <strong>of</strong> the different<br />
surfactant classes extends over a large range, it can be concluded that the<br />
mucous membrane compatibility decreases in the following order:
346 TOXICOLOGY OF SURFACTANTS<br />
nonionic>anionic>cationic surfactants (Draize and Kelley, 1952; Hazleton,<br />
1952; Grant, 1962).<br />
Sensitization<br />
Aside from acute irritation, chemical substances can cause allergies after<br />
contact with the skin or a mucous membrane. The development <strong>of</strong> an<br />
allergy is dependent on certain preconditions. An essential factor is the<br />
individual disposition which is predominantly genetically determined. An<br />
additional important point is the extent <strong>of</strong> damage to the tissue at the point<br />
<strong>of</strong> contact <strong>of</strong> the chemical substance (inflammation), which promotes<br />
sensitization. In addition, the sensitization potential <strong>of</strong> a substance is <strong>of</strong><br />
decisive importance. For products with low molecular weights, this<br />
potential is dependent on their chemical properties. Small molecules are by<br />
themselves not able to trigger a reaction <strong>of</strong> the immune system. They<br />
become immunologically active only after binding to endogeneous<br />
proteins. Since the majority <strong>of</strong> the surfactants can only form weak and<br />
reversible bindings via hydrophobic and electrostatic interactions, this<br />
prerequisite is not fulfilled.<br />
Once the organism is sensitized towards a certain chemical, renewed<br />
contact with trace amounts <strong>of</strong> this material can provoke allergic reactions,<br />
which especially affect the skin and respiratory tract. Typical symptoms are<br />
itching, eczema, exanthema, rhinitis and bronchial asthma.<br />
Anionic surfactants and surfactant containing products were tested for<br />
sensitizing properties by numerous laboratories (Götte, 1967; Kästner,<br />
1980; Siwak et al., 1982) without detecting any significant increase in risk.<br />
The same holds true for nonionic surfactants (Siwak et al., 1982). Some<br />
cationic surfactants, which are able to form stable complexes by the<br />
formation <strong>of</strong> ion pairs with anionic groups <strong>of</strong> proteins, proved to be<br />
allergenic (Schallreuter and Wood, 1986).<br />
Toxicokinetics<br />
Percutaneous absorption<br />
The most important exposure <strong>of</strong> humans occurs through the skin with the<br />
use <strong>of</strong> cosmetics and toiletries. The skin comes in contact with surfactants<br />
also during dishwashing or when washing hands. Since these products are<br />
used over a long period <strong>of</strong> time, possible long-term effects must be<br />
evaluated. Measurement <strong>of</strong> percutaneous absorption <strong>of</strong> surfactants is<br />
important because it provides data for the toxicologist concerning the<br />
amount <strong>of</strong> surfactants which could enter the body through the skin in the<br />
most unfavourable case. Together with other toxicological information,
this allows a realistic evaluation <strong>of</strong> the risk when these compounds are<br />
used.<br />
Due to their economic importance, most studies have been carried out<br />
with anionic surfactants. Fewer studies exist for the other classes <strong>of</strong><br />
surfactants. In vitro measurements <strong>of</strong> the percutaneous absorption <strong>of</strong><br />
sodium dodecyl sulphate indicated a low absorption value for rat skin as<br />
well as for human skin (Blank and Gould, 1961; Embery and Dugard,<br />
1969; Howes, 1975). The low cutaneous absorption <strong>of</strong> sodium dodecyl<br />
sulphate could also be confirmed in experiments with rats (Greb, 1980).<br />
After application <strong>of</strong> a 0.7 per cent aqueous solution <strong>of</strong> sodium dodecyl<br />
sulphate (contact time 15 min), a cutaneous absorption <strong>of</strong> 0.26 µg cm −2<br />
within 24 h was measured (Howes,1975).<br />
In summarizing the results <strong>of</strong> the available studies, one can conclude that<br />
only small amounts <strong>of</strong> surfactants are resorbed through the intact skin.<br />
Since human skin in general is less permeable to chemicals (Rice, 1977;<br />
Wester and Maibach, 1982), the amounts <strong>of</strong> surfactants absorbed<br />
cutaneously in everyday use are probably even smaller. If the epidermis is<br />
removed completely or partially, e.g. damaged skin, the degree <strong>of</strong><br />
absorption can increase substantially (Scala et al., 1968). In vitro studies<br />
demonstrated that cationic surfactants are absorbed by the skin to a much<br />
lesser extent than anionic surfactants (Scala et al., 1968; Geisler, 1976;<br />
Faucher et al., 1979).<br />
The degree <strong>of</strong> percutaneous absorption is generally larger for nonionic<br />
surfactants than for anionic or cationic surfactants. Studies on the<br />
percutaneous absorption <strong>of</strong> alkyl polyethyleneglycol ethers <strong>of</strong> the structure<br />
C 12-(CH 2-CH 2-O) 3H, C 12-(CH 2-CH 2-O) 6H , C 12-(CH 2-CH 2-O) 10H, and<br />
C 15-(CH 2-CH 2-O) 3H were performed under conditions <strong>of</strong> use (Black and<br />
Howes, 1979). The aqueous solutions <strong>of</strong> the applied surfactants were<br />
washed <strong>of</strong>f after a contact time with the skin <strong>of</strong> 15 min. Under these<br />
conditions, the penetration <strong>of</strong> the alkyl polyethyleneglycol ethers was<br />
greater than the penetration <strong>of</strong> the analogous alcohol sulphates or alcohol<br />
ether sulphates. The penetration increased with increasing length <strong>of</strong> the<br />
carbon chain. Percutaneous absorption decreases for an ethylene oxide<br />
content <strong>of</strong> 6 moles or more in the ethoxylate moiety.<br />
Intestinal absorption, metabolism and excretion<br />
W.STERZEL 347<br />
The ingestion <strong>of</strong> surfactants is possible e.g. through the use <strong>of</strong><br />
surfactantcontaining toothpaste, through residues from dishwashing<br />
detergents and through traces <strong>of</strong> surfactants in potable water. Anionic<br />
surfactants are resorbed well in the intestine (Michael, 1968; Black and<br />
Howes, 1980; Bartnik and Künstler, 1987). After absorption, a part <strong>of</strong> this<br />
is excreted together with bile in the faeces and is subject to a enterohepatic<br />
cycle. The majority <strong>of</strong> the absorbed surfactant is metabolized in the liver
348 TOXICOLOGY OF SURFACTANTS<br />
and the respective metabolites are eliminated in the urine. The metabolic<br />
degradation <strong>of</strong> the linear alkyl chain is performed by -oxidation followed<br />
by β-oxidation. The ether-linkage in the ethoxylate portion <strong>of</strong> sulphated<br />
alcohol ethoxylates seems to be resistant to metabolism.<br />
Linear alkylbenzene sulphonates and branched alkylbenzene sulphonates<br />
are metabolized to short chain sulphophenyl carboxylic acids. N-alkyl<br />
sulphates are metabolized by -oxidation <strong>of</strong> the hydrophobic end followed<br />
by β-oxidation. Butyric acid-4-sulphate and acetic acid-2-sulphate are the<br />
end products, which are then further converted in small amounts nonenzymatically<br />
to sulphate and -butyrolactone (Ottery et al., 1970). Studies<br />
by Taylor et al. (1978) demonstrated that alkyl sulphonates are degraded<br />
via the same pathway as alkyl sulphates.<br />
Cationic surfactants can be assumed to be resorbed in the intestine only<br />
to a small extent. This was confirmed in a study with trimethyl cetyl<br />
ammonium bromide (Isomaa, 1975; Isomaa et al., 1976). Due to the low<br />
level <strong>of</strong> resorbed surfactant, an unquestionable identification <strong>of</strong> the<br />
metabolites was not possible. Parts <strong>of</strong> absorbed cationic surfactants were,<br />
as found for anionic surfactants, excreted together with bile in the faeces<br />
and to a lesser degree with the urine.<br />
Nonionic surfactants are resorbed to a large degree in the intestine<br />
(Drotman, 1980). A significant part <strong>of</strong> the material is eliminated with the<br />
bile. Cleavage <strong>of</strong> the ether linkage is obviously possible. Homologous<br />
ethyleneglycol ethers are probably generated as metabolites along with the<br />
corresponding carboxylic acids which are formed through oxidation <strong>of</strong> the<br />
terminal hydroxymethyl group (Drotman, 1980). The sorbitan fatty esters,<br />
which are <strong>of</strong>ten used as emulsifiers, and the ethoxylated fatty acid esters<br />
are hydrolysed in the gastrointestinal tract after oral administration<br />
through cleavage <strong>of</strong> the ester bond. While the resulting fatty acid is treated<br />
metabolically like a natural fatty acid, the polyol component <strong>of</strong> the<br />
sorbitan fatty acid is absorbed in the intestine, but is not further oxidized<br />
and is eliminated predominantly with the urine (Elworthy and Treon, 1967).<br />
Systemic effects<br />
Talking about systemic effects means, in contrast to local effects, the<br />
description <strong>of</strong> reactions arising after the substance has entered the organism<br />
after swallowing, skin penetration or inhalation. For surfactants,<br />
resorption through the skin has to be considered in particular. As described<br />
in the previous section, it is relatively small. But for products that<br />
frequently come into close contact with the skin, either unintentionally or<br />
due to their intended use, the resorption <strong>of</strong> very small amounts over a long<br />
period <strong>of</strong> time cannot be prevented.
Acute toxicity<br />
In general, the acute oral toxicity <strong>of</strong> surfactants is low. The LD 50 values<br />
typically range between several hundred and several thousand mg kg −1 <strong>of</strong><br />
bodyweight. This is <strong>of</strong> the same order <strong>of</strong> magnitude as for table salt<br />
(Swisher, 1968). The most important effects are damage to the mucous<br />
membranes <strong>of</strong> the gastrointestinal tract. High doses induce vomiting and<br />
diarrhea (Weaver and Griffith, 1969). Surfactants exhibit significantly<br />
higher toxicity when the gastrointestinal tract is by-passed through<br />
intravenous injections. Even at very low concentrations, the interaction<br />
with the membrane <strong>of</strong> erythrocytes leads to their destruction. Inhalation <strong>of</strong><br />
surfactant-containing dusts or aerosols in higher concentrations leads to<br />
disturbances <strong>of</strong> the lung function (Coate et al., 1978). This effect can be<br />
attributed to interactions with the surface active film that lines the vesicles<br />
<strong>of</strong> the lung (Kissler et al., 1981). As with local compatibility, there are also<br />
pronounced structure/activity relationships for acute toxicity. Gale (1953)<br />
has investigated the acute toxicity <strong>of</strong> sodium alkyl sulphates from C 8 to C 18<br />
and found the strongest effect for C 12 sulphate.<br />
The anaesthetic properties <strong>of</strong> certain alcohol ethoxylates which can be<br />
observed after intravenous application as well as after application to the<br />
skin or the mucous membranes are remarkable. Ethoxylates <strong>of</strong> unbranched<br />
primary alcohols with 9 ethylene oxide units were found to exhibit local<br />
anaesthetic properties starting with an alkyl chain <strong>of</strong> C 8. The activity<br />
increases with increasing chain length (Zipf and Dittmann, 1964).<br />
Chronic toxicity<br />
W.STERZEL 349<br />
In order to exclude any adverse effects arising from the repeated exposure<br />
against small amounts <strong>of</strong> surfactants over a prolonged period <strong>of</strong> time,<br />
representatives <strong>of</strong> all important classes <strong>of</strong> surfactants were examined for<br />
chronic toxic effects. In these tests, dosages <strong>of</strong> several thousands ppm were<br />
administered over a period <strong>of</strong> up to 2 years. No observable effects were<br />
detected with linear alkylbenzene sulphonates in 2 year studies with rats<br />
using concentrations up to 0.5 per cent (feed) or 0.1 per cent (drinking<br />
water) (Buehler et al., 1971). A sodium alkyl sulphate with an average<br />
chain length <strong>of</strong> C12 was tolerated by rats up to 1 per cent in the feed for 1<br />
year without any remarkable side effects (Fitzhugh and Nelson, 1948). C14 −16α-olefin sulphonates were applied over 2 years in a feeding study in<br />
dosages up to 0.5 per cent without causing any remarkable effect (Hunter<br />
and Benson, 1976). Analogous studies were reported for alcohol<br />
ethoxylates and alkylphenol sulphates, which revealed no toxic symptoms<br />
at doses up to 0.1 per cent and 1.4 per cent, respectively (Larson et al.,<br />
1963; Siwak et al., 1982). Studies on cationic surfactants reported a noobservable-effect-level<br />
<strong>of</strong> 0.25 per cent (Coulston et al., 1961). In all these
350 TOXICOLOGY OF SURFACTANTS<br />
long-term studies, the dosages that were tolerated without damage were in<br />
the range <strong>of</strong> several thousand ppm, indicating large margins <strong>of</strong> safety. This<br />
was confirmed by Hunter and Benson (1976), who calculated for a<br />
relevant example that the respective dosage lies at least by a factor <strong>of</strong> 1000<br />
over the estimated maximum daily exposure level <strong>of</strong> humans. Besides these<br />
data from animal experiments, a series <strong>of</strong> studies exists in which volunteers<br />
ingested considerable amounts <strong>of</strong> anionic or nonionic surfactants over<br />
several weeks, without any noticeable severe adverse effects (Swisher,<br />
1968).<br />
Mutagenicity<br />
Mutagenicity is the induction <strong>of</strong> irreversible changes in genetic material. If<br />
normal cells (somatic cells) are the target, malformation results in<br />
the developing organism. In case <strong>of</strong> the mature organism, it can lead to<br />
tumour formation. If germ cells are the target, the danger exists that the<br />
genetic defect will be passed on to the <strong>of</strong>fspring. All classes <strong>of</strong> surfactants<br />
have been evaluated in numerous test systems. The collected data allow the<br />
conclusion that surfactants pose no considerable risk <strong>of</strong> genetic damage<br />
(Yam et al., 1984; Fowler, 1988; Oba and Takei, 1992).<br />
Carcinogenicity<br />
Due to the widespread use and contact with surfactants the question <strong>of</strong><br />
irreversible damage has to be raised in addition to the problem <strong>of</strong> other<br />
chronic effects. The following compounds were evaluated for<br />
carcinogenicity after administration in the drinking water or feed:<br />
alkylbenzene sulphonate (Buehler et al., 1971), alkyl sulphates (Fitzhugh<br />
and Nelson, 1948), α-olefin<br />
sulphonates (Hunter and Benson, 1976), secalkane<br />
sulphonate (Quack and Rend, 1976), alcohol ether sulphates<br />
(Tusing et al., 1962; Siwak et al., 1982), alcohol ethoxylates (Siwak et al.,<br />
1982) and alkylphenol ethoxylates (Larson et al., 1963; Smyth and<br />
Calandra, 1969). None <strong>of</strong> these experiments provided any indication <strong>of</strong><br />
increased risk <strong>of</strong> cancer after oral ingestion <strong>of</strong> surfactants. The question <strong>of</strong><br />
possible carcinogenic effects <strong>of</strong> surfactants on the skin has also been<br />
studied extensively. Summaries exist by Oba and Takei (1992) and Siwak et<br />
al. (1982).<br />
Embryotoxicity<br />
The effects <strong>of</strong> substances on the organism during pregnancy can lead to<br />
delayed development or death <strong>of</strong> the embryo or malformation. Studies with<br />
the following surfactants revealed no indications <strong>of</strong> embryotoxic activity:<br />
alcohol ethoxylates (Nomura et al., 1980), α-olefin<br />
sulphonates (Palmer et
al., 1975), alcohol ether sulphates and linear alkylbenzene sulphonates<br />
(Nolen et al., 1975). Concerns which started with the publication in 1969<br />
(Mikami et al., 1969) that surfactants had caused malformations in animal<br />
studies could not be reproduced (Oba and Takei, 1980). The findings <strong>of</strong><br />
Mikami et al. (1969) were interpreted to be a result <strong>of</strong> methodical<br />
inadequacies and misinterpretations (Charlesworth, 1976).<br />
Summary<br />
Due to their physico-chemical properties, surfactants are capable <strong>of</strong><br />
reacting with biological membranes, proteins and enzymes. Most <strong>of</strong> their<br />
toxicological properties can be traced back to these interactions. During<br />
the application <strong>of</strong> surfactant-containing products, the most important<br />
aspect <strong>of</strong> consumer safety is local compatibility. No indications <strong>of</strong><br />
systemic, chronic or irreversible damage could be found. Estimates <strong>of</strong> the<br />
amounts <strong>of</strong> orally ingested surfactants typically encountered were reviewed<br />
by several authors. Based on these estimates, a total daily intake <strong>of</strong><br />
surfactants in the range <strong>of</strong> 0.3–3 mg per person was calculated by Swisher<br />
(1968). Due to the low rate <strong>of</strong> percutaneous absorption exposure through<br />
the skin can be neglected. If the above mentioned highest conceivable daily<br />
intake is compared with the dosage that was tolerated without adverse<br />
effects in studies concerning systemic effects, it becomes quite clear that<br />
these amounts can be regarded as harmless. In conclusion, it can be stated<br />
that the use <strong>of</strong> surfactants does not pose a health risk for humans.<br />
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W.STERZEL 351<br />
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354
PART SEVEN<br />
Controversial mechanistic and regulatory<br />
issues in the safety assessment <strong>of</strong><br />
industrial chemicals
25<br />
Low Dose <strong>of</strong> a Genotoxic Carcinogen does not<br />
‘Cause’ Cancer; it Accelerates Spontaneous<br />
Carcinogenesis<br />
WERNER K.LUTZ<br />
University <strong>of</strong> Würzburg, Würzburg<br />
Definitions <strong>of</strong> cancer risk<br />
The risk <strong>of</strong> cancer is normally expressed as a fraction <strong>of</strong> a population<br />
diagnosed with cancer within a specified period <strong>of</strong> time. In an animal<br />
bioassay for carcinogenicity, this period usually is 2 years; in cancer<br />
epidemiology, a life span <strong>of</strong> 65 years (0–64) is <strong>of</strong>ten used. The two periods<br />
can be considered equivalent with respect to the process <strong>of</strong> carcinogenesis:<br />
its rate in different species is inversely correlated with the natural life span<br />
and basal metabolism and the background cancer incidence in 2-year old<br />
rats or mice is very similar to the one seen in 65-year-old humans<br />
(Anisimov, 1989; Raabe, 1989; Tennant, 1993).<br />
For an individual, a cancer risk can only be 0 or 1, depending on<br />
whether the situation is analysed before or after the diagnosis <strong>of</strong> the<br />
tumour. The population-based expression <strong>of</strong> a cancer risk therefore is not<br />
easily visualized and does not take into account interindividual differences<br />
in susceptibility.<br />
In the following discussion, the dose-response relationship in chemical<br />
carcinogenesis is analysed in terms <strong>of</strong> an effect <strong>of</strong> a carcinogen on the<br />
individual tumour latency time (Kodell et al., 1980; Littlefield et al., 1980;<br />
Day, 1983; Gaylor, 1992). Together with the idea <strong>of</strong> background DNA<br />
damage responsible for what is considered ‘spontaneous’ tumour formation<br />
and including individual variability for the rate <strong>of</strong> this process, it will be<br />
shown that low doses <strong>of</strong> genotoxic carcinogens might accelerate the<br />
spontaneous process <strong>of</strong> carcinogenesis but are not expected to induce<br />
cancer ‘out <strong>of</strong> the blue’.<br />
Linear dose response for a DNA-reactive carcinogen<br />
The effect <strong>of</strong> a carcinogen at low dose is normally extrapolated from data<br />
obtained in 2-year bioassays. At the end <strong>of</strong> the 2-year treatment period, the<br />
surviving animals are killed and analysed for the presence <strong>of</strong> tumours. The
fraction <strong>of</strong> tumour-bearing animals is then plotted against the dose, and<br />
lowdose effects are estimated from some model curve fitted to the data<br />
points.<br />
The shape <strong>of</strong> the dose-response curve at the low-dose end is strongly<br />
debated, especially for nongenotoxic carcinogens. For DNA-reactive<br />
carcinogens, it is widely accepted that there is no dose without effect, and a<br />
linear extrapolation is used (Lutz, 1990b). This is based on the idea that 1<br />
molecule <strong>of</strong> a DNA-reactive carcinogen could form a dangerous DNA<br />
adduct in a critical gene and activate an oncogene or inactivate a tumour<br />
suppressor gene by mutation, if the adduct is not repaired before DNA<br />
replication.<br />
Table 25.1 shows the consequences <strong>of</strong> linear interpolation between the<br />
tumour incidence in the controls and in a dosed group <strong>of</strong> a bioassay for<br />
carcinogenicity. An organ-specific tumour incidence <strong>of</strong> 4 per cent in the<br />
control group (2/50) and 14 per cent (7/50) after treatment for 2 years at<br />
10 mg kg −1 per day is assumed. With linear interpolation, a treatmentrelated<br />
increment in tumour incidence <strong>of</strong> 1 per cent would be calculated<br />
per mg kg −1 per day so that at 1 mg kg −1 per day, a 5 per cent total tumour<br />
incidence would be expected.<br />
In humans, increments in cancer risk in the per cent range would not be<br />
acceptable. A risk <strong>of</strong> 1 in 1 million lives might be considered negligible and<br />
the respective exposure could be regarded as ‘virtually safe.’ With the<br />
example given in Table 25.1 and upon linear interpolation, this ‘virtually<br />
safe’ dose would be calculated as 0.0001 mg kg −1 per day.<br />
What does it mean: ‘1 additional tumour in 1000 000<br />
lives’?<br />
The fact that a cancer risk is only 10 −6 cannot be a consolation for the<br />
affected individual. For this person, the cancer risk was 1. In the public<br />
opinion, therefore, an increase by one tumour case per one million lives is<br />
<strong>of</strong>ten interpreted to mean that one additional individual has got cancer<br />
who could otherwise have lived a much longer tumour-free life. For<br />
reasons explained below, this fear appears unfounded.<br />
Endogenous DNA damage; individual susceptibility<br />
W.K.LUTZ 357<br />
Carcinogenesis is a multi-stage process based on the accumulation <strong>of</strong> a<br />
number <strong>of</strong> critical DNA-related changes, due to, for example, DNAcarcinogen<br />
adducts. Evidence <strong>of</strong> background DNA damage from<br />
endogenous and unavoidable substances is accumulating (Ames, 1989;<br />
Loeb, 1989; Lutz, 1990a). It is due to, for instance, electrophiles such as<br />
S-adenosylmethionine, epoxides, quinones, or aldehydes, or to reactive<br />
oxygen species. In addition, DNA is not a chemically stable molecule, it
358 CARCINOGENESIS AT LOW DOSE<br />
Table 25.1 Linear low-dose interpolation <strong>of</strong> a cancer risk based on a hypothetical 2year<br />
bioassay for carcinogenicity<br />
Notes:<br />
Data in boldface; extrapolations in italics.<br />
depurinates and deaminates spontaneously. Finally, DNA replication is not<br />
100 per cent correct so that mutations cannot be avoided completely. The<br />
resulting background DNA damage is responsible for spontaneous<br />
mutations and for what is called spontaneous tumour formation. The<br />
process <strong>of</strong> carcinogenesis therefore has a non-zero rate even if exposure to<br />
exogenous DNA-reactive carcinogens could be avoided.<br />
The level <strong>of</strong> the background mutation rate is expected to show<br />
interindividual variability. It depends both upon genetic and life-style<br />
factors which govern, for instance, enzyme activities responsible for<br />
carcinogen metabolism or DNA repair (Harris, 1989). The rate <strong>of</strong><br />
spontaneous carcinogenesis is further governed by the inherited and<br />
acquired presence <strong>of</strong> activated oncogenes or absence <strong>of</strong> tumour suppressor<br />
genes (Scrable et al., 1990). Therefore, each individual in a heterogeneous<br />
population is expected to have its own endogenous cancer risk expressed as<br />
an individual time-to-tumour or tumourfree lifetime.<br />
Exogenous DNA damage; acceleration <strong>of</strong> spontaneous<br />
carcinogenesis<br />
Exposure to an additional, exogenous DNA-reactive molecule adds to the<br />
background DNA damage, increases the probability <strong>of</strong> a mutation and<br />
accelerates the multi-stage process <strong>of</strong> carcinogenesis. At low doses <strong>of</strong> the<br />
exogenous carcinogen, the rate <strong>of</strong> the process is expected to be dominated<br />
by the background damage so that the exogenous factor cannot constitute<br />
a cancer risk independent <strong>of</strong> the spontaneous process. The acceleration<br />
must be dose dependent and might be related to the background rate<br />
operating in each individual.<br />
It is true, therefore, that even a few molecules <strong>of</strong> a DNA-reactive<br />
carcinogen can have an effect. However, this effect cannot be a tumour
‘out <strong>of</strong> the blue’, in an individual that would otherwise have a low cancer<br />
risk. It ‘only’ reduces the individual’s tumour-free lifetime.<br />
No cancer ‘out <strong>of</strong> the blue’<br />
W.K.LUTZ 359<br />
Figure 25.1 Schematic representation <strong>of</strong> the time course <strong>of</strong> tumour appearance in a<br />
group <strong>of</strong> individuals with large differences in susceptibility. Solid line: background<br />
process <strong>of</strong> spontaneous carcinogenesis; arrows: acceleration <strong>of</strong> the spontaneous<br />
process by exposure to an additional carcinogen.<br />
This interpretation does not contradict the understanding that a low dose <strong>of</strong><br />
a carcinogen could increase the tumour incidence from 40000 to 40001 per<br />
1000000 lives, in the example shown in Table 25.1. The connection<br />
between the two approaches is shown in Figure 25.1. The solid line shows<br />
the appearance <strong>of</strong> a spontaneous tumour in individuals <strong>of</strong> a group <strong>of</strong> 20<br />
people. At the age <strong>of</strong> 65 years, 4 individuals have a tumour diagnosed. This<br />
is equivalent to a cumulative tumour incidence <strong>of</strong> 20 per cent. Exposure <strong>of</strong><br />
this group to an additional exogenous carcinogen would result in some<br />
reduction in the tumour-free lifespan in all individuals. In the example<br />
shown in Figure 25.1, this shift would move one additional individual to<br />
an age <strong>of</strong> diagnosis
360 CARCINOGENESIS AT LOW DOSE<br />
years. This is equivalent to an increase from 40000 to 40001 as a<br />
cumulative incidence 0–64, but it has a completely different meaning. It can<br />
now be excluded that the additional individual would have lived tumour<br />
free for 80 years in the absence <strong>of</strong> the exogenous carcinogen.<br />
Final remarks<br />
With the ideas presented, fear <strong>of</strong> cancer from low dose or from rare<br />
exposures can possibly be reduced. This does not mean that small cancer<br />
risks should be tolerated. Carcinogens in the environment, for instance,<br />
affect all <strong>of</strong> us; the tumour-free life span is reduced in the entire<br />
population.<br />
The model should be valid for tissues with a high spontaneous tumour<br />
incidence and an exponentially steep age dependence indicative <strong>of</strong> a<br />
multistage requirement <strong>of</strong> 5–6 steps. For cells that can be transformed in 2<br />
or 3 steps or that are specifically sensitive in certain phases <strong>of</strong> the<br />
development (in utero or during childhood), the model has to be<br />
reconsidered. This might be necessary for tumours with incidence peaks at<br />
a young age (leukaemia, tumours <strong>of</strong> the lymphatic tissues, brain, testis).<br />
Nevertheless, the latter tumour types are rare in comparison with cancer <strong>of</strong><br />
the old age so that the concept should hold for the majority <strong>of</strong> the tumours<br />
in humans.<br />
References<br />
AMES, B.N., 1989, Endogenous DNA damage as related to cancer and aging,<br />
Mutat. Res., 214, 41–6.<br />
ANISIMOV, V.N., 1989, Dependence <strong>of</strong> susceptibility to carcinogenesis on species<br />
life span, Arch. Geschwulstforsch., 59, 205–13.<br />
DAY, N.E., 1983, Time as a determinant <strong>of</strong> risk in cancer epidemiology: the role <strong>of</strong><br />
multi-stage models, Cancer Surv., 2, 577–93.<br />
GAYLOR, D.W., 1992, Relationship between the shape <strong>of</strong> dose-response curves<br />
and background tumour rates, Regul. Toxicol. Pharmacol., 16, 2–9.<br />
HARRIS, C.C., 1989, Interindividual variation among humans in carcinogen<br />
metabolism, DNA adduct formation and DNA repair, Carcinogenesis, 10,<br />
1563–6.<br />
KODELL, R.L., FARMER, J.H., LITTLEFIELD, N.A., FRITH, C.H., 1980, Analysis<br />
<strong>of</strong> life-shortening effects in female Balb/c mice fed 2-acetylamin<strong>of</strong>luorene, J.<br />
Environ. Pathol. Toxicol, 3 69–88.<br />
LITTLEFIELD, N.A., FARMER, J.H. and GAYLOR, D.W., 1980, Effects <strong>of</strong> dose<br />
and time in a long-term, low-dose carcinogenic study, J. Environ. Pathol.<br />
Toxicol., 3, 17–34.<br />
LOEB, L.A., 1989, Endogenous carcinogenesis: molecular oncology into the<br />
twentyfirst century—Presidential address, Cancer Res., 49, 5489–96.
W.K.LUTZ 361<br />
LUTZ, W.K., 1990a, Endogenous genotoxic agents and processes as a basis <strong>of</strong><br />
spontaneous carcinogenesis, Mutat. Res., 238, 287–95.<br />
LUTZ, W.K., 1990b, Dose response relationship and low dose extrapolation in<br />
chemical carcinogenesis, Carcinogenesis, 11, 1243–7.<br />
RAABE, O.G., 1989, Scaling <strong>of</strong> fatal cancer risks from laboratory animals to man,<br />
Hlth Phys., 57 suppl. 1, 419–32.<br />
SCRABLE, H.J., SAPIENZA, C. and CAVENEE, W.K., 1990, Genetic and<br />
epigenetic losses <strong>of</strong> heterozygosity in cancer predisposition and progression,<br />
Adv. Cancer Res., 54, 25–62.<br />
TENNANT, R.W., 1993, Stratification <strong>of</strong> rodent carcinogenicity bioassay results to<br />
reflect relative human hazard, Mutat. Res., 286, 111–18.
26<br />
Controversial Mechanistic and Regulatory Issues<br />
in Safety Assessment <strong>of</strong> <strong>Industrial</strong> Chemicals —an<br />
Industry Point <strong>of</strong> View<br />
HEINZ-PETER GELBKE<br />
BASF AG, Ludwigshafen<br />
Introduction<br />
In the appropriate classification and risk assessment <strong>of</strong> industrial chemicals<br />
three main players are involved: the scientific community, regulatory<br />
authorities and the chemical industry. The rules <strong>of</strong> the game are given by<br />
the definitions <strong>of</strong> the classification criteria and by guidelines for the risk<br />
assessment process. These definitions and guidelines are sometimes very<br />
flexible and open to different interpretations but in other cases precisely<br />
defined. Mostly these rules have been set up by regulatory bodies, e.g. the<br />
EU or national authorities, but sometimes also by scientific committees like<br />
the MAK commission in Germany or by IARC.<br />
Looking now at these three main players, the scientific community will<br />
provide the data as the starting point for each individual chemical. Possible<br />
controversies centre around the question, how data gaps may be bridged by<br />
scientifically valid assumptions. This indeed may <strong>of</strong>ten be necessary, if a<br />
complete toxicological and mechanistic data base is not available.<br />
Nevertheless, a scientifically based consensus can <strong>of</strong>ten be achieved for this<br />
bridging process.<br />
On the other hand, the other two players—industry and regulatory<br />
authorities—<strong>of</strong>ten do not reach mutually agreed decisions, although both <strong>of</strong><br />
them finally strive for the same target: protection <strong>of</strong> human health and the<br />
environment in an industrialized community. Controversial issues may<br />
sometimes stem from different approaches for bridging the scientific data<br />
gaps, but mostly they arise from political, economic or social aspects,<br />
which <strong>of</strong>ten are not outspoken. Just to give some examples: anticipated<br />
reaction <strong>of</strong> the society, possible emotions <strong>of</strong> the consumer, possible<br />
influences on the next election, availability <strong>of</strong> technical alternatives,<br />
different perceptions <strong>of</strong> the risk-benefit balance, overall economic situation<br />
<strong>of</strong> the community, different evaluations in other countries, impact on<br />
worldwide competitiveness, etc.<br />
In the following an industry point <strong>of</strong> view will be presented for some<br />
specific problems <strong>of</strong> today in the area <strong>of</strong> classification which is a more
qualitative approach and for risk assessment which in addition has to take<br />
into account quantitative aspects. This will be discussed separately for<br />
toxicological effects with thresholds (‘classical’ organ toxicity, reproductive<br />
and developmental toxicity) and without thresholds (mutagenicity,<br />
carcinogenicity). Apart from this subdivision there is one general concern<br />
<strong>of</strong> industry, that is to appropriately take into account exposure.<br />
Exposure<br />
Every classification and especially risk assessment decision should not only<br />
be based qualitatively on the toxicological pr<strong>of</strong>ile, but it should also take<br />
into account quantitatively the toxicological dose-response relationship as<br />
compared to human exposure.<br />
As a first rough approximation the different exposure pr<strong>of</strong>iles may be<br />
grouped into four main categories:<br />
1.<br />
Exposure during chemical production<br />
Many high production volume chemicals are used mainly or even<br />
exclusively as intermediates within the chemical industry. Although large<br />
amounts <strong>of</strong> these materials may be produced or processed within only a few<br />
facilities, exposure is <strong>of</strong>ten quite low and can be controlled or reduced by<br />
technical means. In addition there are many specific features enabling an<br />
efficient exposure control, such as: a trained workforce, site-specific and<br />
personal protection devices, stringent surveillance <strong>of</strong> workforce and work<br />
procedures, medical programmes tailored to the specificities <strong>of</strong> the work<br />
place, well defined exposures at the specific work sites, the exposed<br />
population is well known and limited, specified exposure duration,<br />
relatively homogeneous age and better health status <strong>of</strong> the workforce as<br />
compared to the general population, etc.<br />
2.<br />
Exposure <strong>of</strong> the downstream user during industrial/<br />
manufacturing applications<br />
H.-P.GELBKE 363<br />
In principle, the exposure scenarios may be quite similar to those <strong>of</strong><br />
chemical production, since many <strong>of</strong> the features described above also relate<br />
to or may be implemented at smaller workshops <strong>of</strong> the downstream user.<br />
Unfortunately, in reality <strong>of</strong>ten quite high exposures prevail in small<br />
workshops, possibly due to limited expenditures into exposure reduction<br />
measures or to a workforce not specifically trained for handling <strong>of</strong><br />
dangerous chemicals.
364 CONTROVERSIAL ISSUES IN SAFETY ASSESSMENT<br />
3.<br />
Exposure <strong>of</strong> the consumer<br />
When looking at the vast array <strong>of</strong> chemicals in use today, most <strong>of</strong> them<br />
will have industrial applications and only relatively few will directly be<br />
used by consumers. Especially highly reactive chemicals with their inherent<br />
potential for health hazards will generally not enter into consumer<br />
application fields simply due to their limited stability. But if a chemical gets<br />
to the consumer, efficient control measures can hardly be implemented, be<br />
it for the exposure per se, the exposed population, appropriate handling or<br />
prevention <strong>of</strong> misuse.<br />
4.<br />
Exposure <strong>of</strong> the general population via the environment<br />
Apart from highly reactive substances, all chemicals will to some extent<br />
enter the environment depending on their application fields and their<br />
processing to other end products. The environmental concentrations are<br />
determined by the amount released, by the distribution media, local<br />
situations and the efficacy <strong>of</strong> the different degradation processes. In<br />
general, the exposure <strong>of</strong> the population via the environment will be very<br />
low as compared to the workplace and health hazards are not to be<br />
expected. Of course highly persistent and highly toxic substances can be an<br />
exemption to this rule and should be monitored carefully.<br />
These exposure scenarios can be exemplified by a textile dyestuff:<br />
manufacturing within the chemical industry makes use <strong>of</strong> various starting<br />
materials and intermediates, which will not end up—apart from minute<br />
impurities—in the final product, and downstream exposure to these<br />
materials will be negligible. The manufacture <strong>of</strong> the dyestuff, its further<br />
handling, processing and formulation within the chemical industry can<br />
easily be controlled. But when this material is used downstream in textiledyeing<br />
workshops an efficient exposure control and appropriate handling<br />
may not always be guaranteed and higher exposures are conceivable.<br />
During the dyeing process, parts <strong>of</strong> the material enter into the<br />
environment, for example into aqueous media. This might lead to an<br />
exposure <strong>of</strong> the general population, albeit at very low concentrations.<br />
Consumer exposure can occur via migration <strong>of</strong> the dyestuff from the<br />
textile, sweat being the carrier medium or for small children the saliva, but<br />
exposure will most <strong>of</strong>ten be so low that a health hazard is not to be<br />
expected.<br />
It is one <strong>of</strong> the main concerns <strong>of</strong> industry within the processes <strong>of</strong><br />
classification and risk assessment that sufficient consideration is <strong>of</strong>ten not<br />
given to the different exposure pr<strong>of</strong>iles <strong>of</strong> each chemical. This results in<br />
simplified black and white decisions, which are not very helpful for an
appropriate and cost effective health protection within our industrialized<br />
world.<br />
Threshold effects<br />
Classification<br />
H.-P.GELBKE 365<br />
The classification system <strong>of</strong> the EU resides in the designation <strong>of</strong><br />
appropriate R-phrases. In most cases they are rather meant for hazard<br />
identification (e.g. irritation, sensitization, carcinogenicity) and not so<br />
much for risk characterization. Sometimes, risk aspects are also involved<br />
when specific dose levels are decisive for a specific R-phrase (e.g. acute<br />
toxicity). Thereby not only the effect per se is taken into account but also<br />
the strength <strong>of</strong> the effect.<br />
This latter principle also applies to the R 48-phrase (‘danger <strong>of</strong> serious<br />
damage to health by prolonged exposure'), if severe (irreversible) toxic<br />
effects are observed at dose levels <strong>of</strong> ≤50 mg kg −1 body weight per day in a<br />
90-day test after oral administration. For other routes and durations <strong>of</strong><br />
exposure similar provisions exist. The strategy to use dose limits is<br />
certainly appropriate for threshold effects.<br />
Although for reproductive and developmental effects thresholds are also<br />
accepted in most cases, no such dose limits are given apart from 1000 mg<br />
kg −1 body weight per day. This dose limit is not equivalent to that for the R<br />
48-phrase, because it only stems from the limit dose levels <strong>of</strong> the test<br />
guideline and is higher by more than one magnitude. Thus, for<br />
reproductive/ developmental toxicity classification, exposure and risk<br />
considerations are not influential in contrast to the R 48. Such a simplified<br />
‘yes or no’ approach for a threshold effect is not justifiable neither from a<br />
scientific point <strong>of</strong> view nor for an appropriate health protection. This is<br />
even more so, if the proposal for the ‘Restrictions on Marketing and Use<br />
Directive’ (13th Amendment to Directive 76/769/EEC) is implemented<br />
calling for a general prohibition <strong>of</strong> category 1 and 2 reproductive/<br />
developmental toxicants in consumer products in concentrations <strong>of</strong> ≥0.5%<br />
—if no specific concentration limit has been accepted according to the<br />
preparation guideline (Commission Directive 93/18/EEC; Council Directive<br />
88/379/EEC).<br />
The inconsistency <strong>of</strong> the approaches for ‘classical’ organ toxicity and<br />
developmental/reproductive toxicity can easily be demonstrated by the<br />
following theoretical example:<br />
For neurotoxicity a LOEL <strong>of</strong> 40 mg kg −1 day −1 in a 90-day oral test will<br />
result in a R 48-phrase, but a LOEL <strong>of</strong> 80 mg kg −1 day −1 would not<br />
lead to classification. On the other hand, slight foetal weight
366 CONTROVERSIAL ISSUES IN SAFETY ASSESSMENT<br />
reduction or impairment <strong>of</strong> fertility observed at 800 mg kg −1 day −1 in<br />
anappropriate developmental or reproductive test without parental<br />
toxicity would lead to classification into the respective category 3 or<br />
possibly category 2, even if a NOEL was found at 400 mg kg −1 day −1 .<br />
This certainly is not appropriate when considering the reversibility<br />
and severity <strong>of</strong> the effects and the differences in the NOELs and<br />
LOELs <strong>of</strong> one order <strong>of</strong> magnitude.<br />
Risk assessment<br />
The general strategy in risk assessment <strong>of</strong> chemicals with threshold effects<br />
is to use ‘assessment’ factors (‘uncertainty’ or ‘safety’ factors) (AF) for<br />
setting appropriate exposure limits. This concept was originally introduced<br />
by the WHO to establish ADI values (acceptable daily intake) for pesticide<br />
residues in food. Here generally a ‘safety’ factor <strong>of</strong> 100 is applied to the<br />
NOEL <strong>of</strong> a chronic experiment; higher or lower factors might be used for<br />
specific effects or experimental conditions.<br />
This basic approach can be generalized from consumer exposure to<br />
pesticide residues in food to other chemicals and exposure scenarios. The<br />
main problem then will be the selection <strong>of</strong> an appropriate AF. First <strong>of</strong> all,<br />
there are some general considerations to be taken into account:<br />
– Should AFs for the workforce and the general population differ because<br />
<strong>of</strong> the age characteristics and the general health status <strong>of</strong> workers?<br />
– Should different AFs be selected for chemicals and pesticides, taking into<br />
account that pesticides are specifically tailored for biological activity?<br />
– What are suitable AFs for route to route extrapolations if the<br />
experimental exposure does not correspond to that <strong>of</strong> humans? Thereby<br />
metabolic firstpass effects in the liver and different efficacies <strong>of</strong> the<br />
adsorption barriers <strong>of</strong> skin, lung and the intestines have to be taken into<br />
account.<br />
– What is the appropriate dose parameter, mg kg −1 body weight, mg m −2<br />
surface area or concentration?<br />
– What are suitable AFs for developmental effects which may occur in<br />
principle after a single exposure and may lead to irreversible lifetime<br />
impairment?<br />
– Are specific AFs necessary for toxic effects on the reproductive organs as<br />
compared to toxic effects on other organ systems?<br />
Apart from these general considerations there are also specific criteria<br />
decisive for the selection <strong>of</strong> AFs depending on each single chemical, its<br />
total data base and the experimental details. Very <strong>of</strong>ten the final AF is<br />
obtained by additional default factors which are to account for
H.-P.GELBKE 367<br />
experimental insufficiencies or to bridge data gaps. Just to give some<br />
examples for such specific considerations:<br />
– reversible versus irreversible effects,<br />
– duration <strong>of</strong> the study,<br />
– NOAEL versus NOEL,<br />
– LOAEL versus NOAEL,<br />
– local versus systemic effects,<br />
– species-specific effects,<br />
– species differences in anatomy or physiology,<br />
– similar versus different results observed in experiments with various<br />
species,<br />
– biokinetics and metabolism (e.g. metabolic pathways are species specific<br />
or occur only at high doses),<br />
– structure-activity considerations.<br />
This listing certainly not being complete clearly demonstrates that<br />
appropriate AFs cannot be arrived at by a simple cook-book procedure,<br />
but a flexible case-by-case approach is required for each individual<br />
chemical and data set. This is extremely important in order to avoid overconservative<br />
AFs; and in the long run over-conservative risk assessments<br />
are just as prohibitive for an appropriate health protection in an<br />
industrialized world as an underestimation <strong>of</strong> risk may lead to a more<br />
immediate danger to health. Over-conservative risk evaluations will result<br />
in a wrong allocation <strong>of</strong> resources, an unjustified prohibition <strong>of</strong> valuable<br />
chemicals, a wrong or unnecessary selection <strong>of</strong> alternative materials. etc.<br />
What might be an indication for an over-conservative AF? In principle,<br />
AFs for threshold effects should then be questioned to be over-conservative<br />
if they lead to acceptable human exposures which would also be<br />
appropriate for non-threshold effects (e.g. carcinogenicity) This can be<br />
exemplified by the following consideration:<br />
For a carcinogenicity experiment a ‘LOAEL’ in classical terms would<br />
be equivalent roughly to a dose just leading to a statistically increased<br />
tumour incidence <strong>of</strong> about 5 per cent. A ‘virtual NOAEL’ in the same<br />
classical sense without a statistically significant increase could then be<br />
at a dose with an actual tumour incidence <strong>of</strong> 1 per cent, which will<br />
not show up as a substance related effect under usual experimental<br />
conditions. At such a dose level the extra tumour risk would be 1/100.<br />
Applying an AF <strong>of</strong> 1000 to this ‘virtual NOAEL' would result in an<br />
exposure level with a risk <strong>of</strong> 1/10 5 , and an AF <strong>of</strong> 10000 in one with a<br />
risk <strong>of</strong> 1/10 6 using a simple linear extrapolation without further<br />
default considerations. Exposure levels with a risk <strong>of</strong> 1/10 5 or 1/10 6 are<br />
under discussion as ‘virtually safe doses’ for the workforce or the
368 CONTROVERSIAL ISSUES IN SAFETY ASSESSMENT<br />
general population. Thus, AFs <strong>of</strong> >1000 should always be questioned<br />
as possibly over-conservative for threshold effects, since exposure<br />
levels thereby obtained could also be acceptable for non-threshold<br />
effects like carcinogenicity. The assumptions underlying such high AFs<br />
should be re-examined critically.<br />
In the light <strong>of</strong> these considerations AFs for a developmental toxicity <strong>of</strong><br />
1000– 5000, as they are under discussion by some groups today, should<br />
also be questioned as possibly being over-conservative. It should not only be<br />
taken into account that developmental effects most <strong>of</strong>ten will have<br />
thresholds but also that their severities span a wide range from slight foetal<br />
weight impairment up to disabling malformations.<br />
In addition an unreflected selection <strong>of</strong> default factors in ‘classical’ organ<br />
toxicity can easily lead to over-conservative AFs: starting with a factor <strong>of</strong><br />
10 each for inter- and intra-species variability yields the AF <strong>of</strong> 100 used for<br />
ADI-calculations. In addition the following default factors are sometimes<br />
proposed:<br />
– Extrapolation from subacute/subchronic exposure to chronic exposure:<br />
a default factor <strong>of</strong> 10.<br />
– Extrapolation to the NOEL, if only a LOEL was obtained: a default<br />
factor <strong>of</strong> 2–5.<br />
– Taking into account an inappropriate experimental design: a default<br />
factor <strong>of</strong> 2–5–10.<br />
Thereby, for a multiple dose study with a marginal effect at the lowest dose<br />
level and an experimental design not fully in accordance with today’s<br />
standards, these default factors would result in a final assessment factor <strong>of</strong><br />
4000– 50000. It is highly questionable whether such an AF is really<br />
appropriate for threshold effects in comparison to the example given above<br />
for carcinogenicity.<br />
Non-threshold effects<br />
Other principles and approaches for classification and risk assessment have<br />
to be applied for non-threshold effects since safe exposure levels cannot be<br />
defined at which an adverse health effect will definitely not occur. Thus,<br />
for these compounds carcinogenic and mutagenic effects cannot be<br />
excluded even at very low dose levels albeit with extremely low<br />
probability.
Classification<br />
H.-P.GELBKE 369<br />
The present classification systems <strong>of</strong> scientific organizations (e.g. IARC,<br />
German MAK-commission) or regulatory agencies (e.g. EPA, EU<br />
commission) reside in quite a simplistic ‘strength <strong>of</strong> evidence’ approach:<br />
how valid are the experimental or epidemiological data?<br />
In future we should strive for a ‘weight <strong>of</strong> evidence’ approach which<br />
appears much more appropriate for a realistic human health protection,<br />
since it takes into consideration both risks and benefits <strong>of</strong> man-made and<br />
natural chemicals. Such a classification system basically asks the question:<br />
what do the experimental data really mean to humans at specific exposure<br />
levels? Thereby both qualitative and quantitative aspects are considered.<br />
Qualitatively whether and to what extent the mechanisms leading to an<br />
adverse effect in animals will also act in humans, and quantitatively to put<br />
the experimental dose-response relationship into context with human<br />
exposure. Of course, worst case exposure scenarios have to be taken into<br />
account. If under these considerations the experimental carcinogenic effect<br />
would not be relevant for humans a classification would not be<br />
appropriate.<br />
This latter quantitative aspect has led to discussions in several groups, as<br />
to whether a separate category for ‘weak carcinogens’ should be<br />
established, since more and more compounds turn out to be experimental<br />
carcinogens with a very weak or questionable effect. Such a category could<br />
be used for example for compounds which:<br />
– would only have insignificant effects even under worst case exposure<br />
scenarios,<br />
– did not give a carcinogenic response in appropriate animal experiments<br />
but are metabolized to carcinogenic intermediates or exert genotoxic<br />
effects in vivo,<br />
– show metabolic toxification to genotoxic metabolites only at high doses<br />
where the ‘normal’ metabolic detoxification pathway is overwhelmed.<br />
It is doubtful whether such a new category would really mean a step<br />
forward and be helpful. First <strong>of</strong> all why should compounds be classified if<br />
the carcinogenic effect is not to be expected in humans even under worst<br />
case exposure scenarios? And secondly how can the message <strong>of</strong> ‘weak<br />
carcinogenicity’ be brought over to the public without raising an emotional<br />
over-reaction to these compounds.<br />
Apart from these considerations on the general approach (‘strength’ or<br />
‘weight <strong>of</strong> evidence’) there is one specific major problem: presently, only<br />
criteria for classification are well defined, but those for non-classification<br />
are either not or only very vaguely described. It is also an important<br />
challenge to set up practicable and clear-cut non-classification criteria
370 CONTROVERSIAL ISSUES IN SAFETY ASSESSMENT<br />
which can be used by industry to tailor experiments in order to refute the<br />
classification <strong>of</strong> a questionable animal carcinogen. In the long run there<br />
will be no benefit if about 50 per cent <strong>of</strong> all chemicals are classified as<br />
carcinogens, neither for the public nor for industry nor for an adequate<br />
protection <strong>of</strong> human health.<br />
Risk assessment<br />
Most scientific committees and regulatory agencies refrain from<br />
scientifically based risk assessments for carcinogens but rather propagate<br />
an exposure as low as possible. And if a risk assessment is really carried<br />
out, it usually just applies a simplistic mathematical extrapolation using the<br />
linearized multistage model and highly conservative default assumptions to<br />
bridge data gaps. These mathematical procedures arrive at a scientifically<br />
unjustified numerical precision <strong>of</strong> the risk estimate. One <strong>of</strong> the problems is<br />
to explain to the public the real meaning and the uncertainties <strong>of</strong> such a<br />
risk assessment. A possible alternative could be to substitute the<br />
mathematical extrapolation by an appropriate assessment factor which <strong>of</strong><br />
course has to take into account the severity and irreversibility <strong>of</strong> the<br />
carcinogenic effect. The simplistic mathematical modelling might be used<br />
only for selection <strong>of</strong> priority chemicals for further in-depth investigations.<br />
On the other hand, a mathematical risk assessment can be an<br />
appropriate procedure for chemicals with a broad experimental data base,<br />
when the most relevant default assumptions are substituted by real data.<br />
This would be <strong>of</strong> primary importance for:<br />
– the selection <strong>of</strong> the mathematical model: the simplistic linearized<br />
multistage model presently in use could be substituted by biologically<br />
driven models, like that proposed by Sielken (1989) or the MVK-model<br />
(Moolgavkar and Knudson, 1981; Moolgavkar et al., 1988).<br />
– dose scaling from animals to humans: presently the experimental dose in<br />
mg kg −1 body weight is <strong>of</strong>ten extrapolated to humans by transforming<br />
the dose to mg m −2 body surface. This is used both for compounds<br />
which are metabolically toxified and detoxified, the scientific basis for<br />
such an undifferentiated procedure is at best highly doubtful. In the<br />
future this default assumption could be substituted by physiologically<br />
based pharmacokinetic (PBPK) modelling.<br />
– estimation <strong>of</strong> the target dose: presently the external dose to which the<br />
animals are exposed is considered to be proportional to the dose<br />
reaching the target tissue or the target chemical entities—generally the<br />
DNA. Again in future this could be substituted by adequate PBPK<br />
modelling.
For the time being, there are only few compounds with an experimental<br />
data base broad enough to substitute these default assumptions in every<br />
respect. But there are important industrial chemicals—and their number<br />
will increase eventually—for which at least some data are available for a<br />
justified substitution <strong>of</strong> part <strong>of</strong> the default assumptions. And this should be<br />
done as far as possible when a mathematical risk assessment is carried out.<br />
Conclusions<br />
Problems <strong>of</strong> classification and risk assessment have been discussed<br />
separately for toxicological effects with thresholds (‘classical’ organ<br />
toxicity, reproductive and developmental toxicity) and without thresholds<br />
(mutagenicity, carcinogenicity). With regard to the general procedures <strong>of</strong><br />
today, for industry there are two main points <strong>of</strong> concern:<br />
1. Not only for risk assessment but also for classification, exposure<br />
considerations should be taken into account. In principle, there are<br />
four different exposure scenarios: (a) for chemical production or within<br />
chemical industry, (b) for industrial application by downstream users,<br />
(c) for the consumer and (d) for the general population via the<br />
environment. For a given chemical the exposure may vary widely for<br />
the different scenarios, and there are many chemicals which are only<br />
used within chemical industry, which will never reach the consumer or<br />
which will enter into the environment only in minute quantities.<br />
Exposure estimates, including worst case scenarios, should be included<br />
in the processes <strong>of</strong> risk assessment and classification in order to avoid<br />
over-conservative results, which are not in the interest <strong>of</strong> adequate<br />
health protection.<br />
2. To get away from risk assessment procedures based on default<br />
assumptions which will <strong>of</strong>ten lead to over-conservative results; a case<br />
by case approach making use <strong>of</strong> all available data is scientifically far<br />
more appropriate.<br />
These problems have been elaborated for the:<br />
H.-P.GELBKE 371<br />
– classification <strong>of</strong> chemicals for reproductive or developmental toxicity<br />
within the regulatory framework <strong>of</strong> the EU,<br />
– selection <strong>of</strong> appropriate ‘assessment factors’ for chemicals with<br />
threshold effects,<br />
– classification and risk assessment <strong>of</strong> chemicals with non-threshold<br />
effects (especially carcinogens).
372 CONTROVERSIAL ISSUES IN SAFETY ASSESSMENT<br />
References<br />
MOOLGAVKAR, S.H. and KNUDSON, A.G., 1981, Mutation and cancer: a<br />
model for human carcinogenesis, Journal <strong>of</strong> the National Cancer Institute, 66,<br />
1037–52.<br />
MOOLGAVKAR, S.H., DEWANJI, A. and VENZON, D.J., 1988, A stochastic<br />
two-stage model for cancer risk assessment, I. The hazard function and the<br />
probability <strong>of</strong> tumour. Risk Analysis, 8, 383–92.<br />
SIELKEN, R.L.JR, 1989, Useful tools for evaluating and presenting more science in<br />
quantitative cancer risk assessments, Toxic Substances Journal, 9, 353–404.
Absorption <strong>of</strong> organic solvents 3–7<br />
Acceptable daily intake (ADI) 43, 168<br />
Acetylcholinesterase (AChE) 238<br />
N-acetyl-S-(5-ethoxy-l,2,4-thiadiazol-3yl-methyl)<br />
-L-cysteine (ET-MA) 26<br />
N-acetyl-β-glucosaminidase (NAG) 117<br />
Acid anhydrides 152–7<br />
Acrylamide 238<br />
Acrylonitrile 23, 313<br />
Adduct determination 81–7<br />
limitations <strong>of</strong> methods for 84<br />
Adduct formation <strong>of</strong> reactive<br />
compounds 74<br />
Adipate esters 223<br />
Airway hyperreactivity 128, 131<br />
Airway inflammation, animal models<br />
129<br />
Alachlor 208<br />
Alcohol ethoxylates, anaesthetic<br />
properties 349<br />
Alkaline phosphatase (ALP) 117<br />
Alkylating agents 76–9<br />
Alkyldeoxyguanosine adducts 186<br />
Allergic contact dermatitis 137<br />
Allometric equation 44, 46<br />
Allometric scaling 44–55<br />
Allometry 44<br />
Antioxidants 316–39<br />
Aromatic amines 74, 77<br />
Ascorbate 241<br />
Assessment factors (AFs) 366–71, 370<br />
Astrocytes 240<br />
ATP 78, 79, 159<br />
Axonal proteins 238<br />
Azo colorants 301<br />
carcinogenic 301–6<br />
Index<br />
Benzene 39<br />
Benzidene 301<br />
Benzo(a)pyrene (BP) 159, 186<br />
1,4-benzothiazines 64<br />
Benzotriazole-based light stabilisers<br />
322–29<br />
blood kinetics 323–29<br />
blood metabolites 323–29<br />
effects on rat dam and foetal liver<br />
329–4<br />
in vitro hydrolysis 323<br />
liver enzyme induction 327–31<br />
safety assessment 331, 334–7<br />
BGA, collaborative study 199<br />
Bioactivation mechanisms 11, 35–41<br />
Biocides 208<br />
Bioinactivation mechanisms 11<br />
Biological effect monitoring (BEM) 19<br />
Biological effects, determination <strong>of</strong> 189–<br />
6<br />
Biological monitoring (BM) 19–1<br />
definition 2<br />
organic solvents 2–3<br />
Biomarkers<br />
glutathione conjugation products as<br />
20–2<br />
<strong>of</strong> neurotoxicity 238–4<br />
Biotransformation 12<br />
Bisphenol A diglycidylether (BPADGE)<br />
83<br />
Body metabolic potential 48<br />
Body surface area (BSA) 46–8<br />
BP-DNA adducts 187<br />
Bromobenzene 63<br />
Bromo-diglutathionyl hydroquinones<br />
64
374 INDEX<br />
2-bromo-glutathionyl 64<br />
Bromohydroquinone 63–5<br />
o-bromophenol 63<br />
Bronchial challenge tests 149<br />
Bronchial hyperreactivity 151<br />
Bronchoalveolar lavage 117, 123<br />
1,3-butadiene<br />
PBPK/PBTK model 16<br />
physiologically based toxicokinetic<br />
modeling 31–3<br />
2-butoxyacetic acid (BAA) 172–80<br />
2-butoxyethanol, PBPK/PBTK model<br />
166–4, 172–80<br />
Calcium 245–9, 250, 267<br />
Cancer risk<br />
‘absolute’ 191<br />
assessment 188–6<br />
definitions 356<br />
management <strong>of</strong> 181<br />
Carbamazepine 322<br />
Carbendazim 285–8<br />
Carbon disulphide 238<br />
Carcinogen(esis) 179–6, 356<br />
azo colorants 301–6<br />
DNA-reactive 356<br />
dose-response relationship in 356<br />
epidemiological approaches to<br />
detect and identify 183–9<br />
genotoxic 181–8<br />
identification <strong>of</strong> 183–93<br />
peroxisome proliferation 224–30<br />
role <strong>of</strong> CYPs in activation and<br />
detoxication <strong>of</strong> 206<br />
spontaneous process 356, 357–3<br />
surfactants 350<br />
Carcinogenic potency 181<br />
Carcinogenic Substances Regulation<br />
301<br />
Catabolic metabolism 260<br />
Catalase 241<br />
Cell-mediated immune responses in<br />
chemical respiratory allergy 142–6<br />
Cellular nucleophiles 73<br />
Cerebral calcium accumulation 240<br />
Chemical pesticides, immunotoxicity <strong>of</strong><br />
203<br />
Chemical respiratory allergy, cellmediated<br />
immune responses in 142–6<br />
Chlordecone 241<br />
Chlorinated solvents 223<br />
Chlor<strong>of</strong>orm 38–1<br />
Chromates 301<br />
Chrysotile asbestos fibres<br />
pulmonary toxic effects 116–30<br />
size-separation methods 118–3<br />
Classification systems 368–3<br />
Color Index 301<br />
Colorants (dyes and pigments) 301–10<br />
regulatory aspects (FRG) 304–7<br />
Cosmetics 208<br />
Cultured porcine thyrocytes 261<br />
Cyclophosphamide 286–90<br />
Cytochrome P-450 (CYP) 15, 60, 74,<br />
206–15, 318, 327, 332<br />
Cytokine products <strong>of</strong> murine Th 1 and<br />
Th 2 cells 141<br />
Cytokines 140<br />
Dangerous compounds 208<br />
1,2-DCV-Cys 29, 31<br />
DCV-G 29<br />
1,2-DCV-G 29, 31<br />
1,2-DCV-Nac 29<br />
2,2-DCV-Nac 29<br />
Dermal uptake <strong>of</strong> organic solvents 5–7<br />
Detoxication, enzymes involved in 60–4<br />
Developmental neurotoxins 247–2<br />
Diagonal radioactive zones (DRZ) 158<br />
Diarylide pigments 81–3<br />
1,2-dibromoethane 40, 65–8<br />
3,3′-dichlorobenzidine (DCB) 81–3,<br />
301, 303<br />
1,2-dichloroethane 65–8<br />
2,5-dichloro-3-(glutathion-S-y1)<br />
hydroquinone 64<br />
Dichloromethane 39<br />
1,3-dichloropropene (DCP) 21–3<br />
S-(l,2-dichlorovinyl)glutathione. See 1,<br />
2-DCV-G<br />
di-(2-ethylhexyl)adipate (DEHA) 223,<br />
224, 227, 229–4<br />
di-(2-ethylhexyl)phthalate (DEHP) 223,<br />
224, 227–4, 286, 287
5α-dihydrotestosterone 212<br />
Diisocyanate asthma 151<br />
di-(isodecyl)phthalate 223<br />
3,3′-dimethoxybenzine 301<br />
3,3′-dimethylbenzidine 301<br />
Dioctyl phthalate 152<br />
Diphenylhydantoin 322<br />
2,6-di-tert-butyl-4-methyl phenol<br />
(BHT) 316–20<br />
DNA adducts 74, 77–81, 83, 84, 158,<br />
159–9, 183–91, 189, 224<br />
immunoenrichment <strong>of</strong> 185–2<br />
DNA binding 60<br />
DNA damage 183, 192, 208, 356<br />
endogenous 356–1<br />
exogenous 357–3<br />
DNA reactions 73<br />
DNA-reactive carcinogens 356<br />
dose-response curve 356<br />
DNA repair 192<br />
DNA replication 180, 192, 357<br />
DNA synthesis 227, 230<br />
Dopamine 246<br />
Dose determination 188<br />
Dose-response relationship<br />
and exposure pr<strong>of</strong>iles 362–68<br />
DNA-reactive carcinogens 356<br />
in carcinogenesis 356<br />
low-dose range 190<br />
DTH reactions 196<br />
Dyes. See Colorants (dyes and<br />
pigments)<br />
EC annex VII and VIII toxicity tests<br />
280<br />
ECETOC 202<br />
EDB, conjugation in rats and man 67<br />
Electrophilic agents 60, 71<br />
Electrophilic centres 72<br />
Electrophilic compounds, examples <strong>of</strong><br />
72<br />
Electrophilic metabolites 60<br />
Embryotoxicity tests 289–3<br />
Emulsifiers 312<br />
Endocrine dysfunction, xenobioticinduced<br />
254<br />
INDEX 375<br />
Endocrine toxicity, classification <strong>of</strong> 254–<br />
59<br />
Endocrine toxicology, thyroid 254–80<br />
Entire mammalian tests 283<br />
Environmental exposure 363<br />
Environmental monitoring (EM) 19–1<br />
Epoxide hydrolases (EH) 60<br />
Equivalent radiation dose concept 192–<br />
8<br />
Ethanol 241<br />
5-ethoxy-1,2,4-thiadiazole-3-carboxylic<br />
acid (ET-CA) 25–7<br />
Ethylene glycol methyl ether (EGME)<br />
286–91<br />
Etridiazol, disposition <strong>of</strong> 25–7<br />
European Union (EU) 167<br />
Exposure pr<strong>of</strong>iles 362–68<br />
Fatty amines 312<br />
FDA Segment I study for medicines 283<br />
Fecundity tests 284<br />
Fertility<br />
and embryotoxicity 285<br />
toxicity 281<br />
Fibre aerosols 97<br />
Fibre glass 91, 99–4<br />
airborne levels in workplace 109<br />
industrial hygiene studies 105–10<br />
lung fibre levels in workers 110<br />
Fibre recovery from lung tissue 118<br />
Fischer 344 study (Kimber-White) 198–<br />
4<br />
Flame retardant 312–18<br />
Flavin-containing monooxygenase<br />
(FMO) 211–16<br />
Fluoranthene-DNA adducts 187<br />
Foetal abnormalities 291<br />
Food additives 208<br />
Food and Agriculture Organization 208<br />
Fotemustine 46, 47<br />
Free radical formation 240–4, 246<br />
Full scale testing 288–1<br />
Furazolidone 68<br />
Gas chromatography (GC) 74, 76, 81
376 INDEX<br />
Gas chromatography/mass<br />
spectrometry (GC/MS) 74–9, 81, 84,<br />
185<br />
Genotoxic carcinogens 168<br />
Genotoxic hazards 181–8<br />
Genotoxicity 184, 281<br />
in vitro assays 184<br />
Germ cell mutagens 168<br />
Glial fibrillary acidic protein (GFAP)<br />
240–7, 247–2<br />
Gliotypic proteins 238–3<br />
Glutathione (GSH) 15, 20, 24, 184,<br />
241<br />
Glutathione conjugation, reversible 67–<br />
9<br />
Glutathione S-transferase (GST) 15, 20,<br />
24, 60–4, 66, 161, 212, 329<br />
Glycophorin A 160<br />
GM-CSF 143<br />
Growth desensitising mechanism<br />
(GDM) 263, 265<br />
Gunn rat hepatocytes in vitro, studies<br />
on 274<br />
Haemoglobin 76<br />
Haemoglobin adducts 189<br />
Health surveillance (HS) 19<br />
Heavy metals 244–9<br />
Hepatic metabolism, xenobiotics acting<br />
on 267–76<br />
‘Hepatic pharmacokinetic stuff’ 49<br />
Hepatocarcinogenesis, mechanisms <strong>of</strong><br />
226–1<br />
Hepatocytes<br />
co-cultures<br />
phase 1 reactions 210–16<br />
phase 2 reactions 212–17<br />
long-term cultures <strong>of</strong> 207–14<br />
Herbicides 208<br />
Hexamethylene diisocyanate (HDI) 151<br />
n-hexane 3, 7, 238<br />
2,5-hexanedione 3<br />
HGPRT mutation assay 192<br />
HHPA 152<br />
Himic anhydride (HA) 152<br />
Hormone elimination 269<br />
Hormone synthesis 258<br />
Host resistance (HR) studies 200<br />
HPLC 74–83, 84<br />
Human allergic disease 142<br />
Human clearance prediction 49, 53<br />
Human exposure monitoring 188<br />
Human unbound clearances 51–3<br />
Hydroxymethylethenodeoxyadenosine<br />
(HMEdA) 83<br />
1-hydroxypyrene 159<br />
Hypersensitivity reactions, Type I-IV<br />
196<br />
Hypothalamic-pituitary-thyroid-liver<br />
(H-P-T-L) axis 256–60<br />
investigative tests on 267<br />
thyroid toxicity via 265–76<br />
toxicological 266<br />
xenobiotic toxic effects on 258<br />
Hypoxanthin guanine phosphoribosyl<br />
transferase (HPRT) 160<br />
ICICIS collaborative study 197–3<br />
IFN- 140–6<br />
IgA 199, 200<br />
IgE 153<br />
IgE antibody 138<br />
IgE antibody responses, induction and<br />
regulation <strong>of</strong> 140–5<br />
IgG 153, 200<br />
IgG2a antibody 140<br />
IgM 200<br />
Immune system, evaluation <strong>of</strong> toxicity<br />
to 196–10<br />
Immunoenrichment <strong>of</strong> DNA adducts<br />
185–2<br />
Immunological methods 77, 78<br />
Immunotoxic side effects, screening <strong>of</strong><br />
197<br />
Immunotoxicity <strong>of</strong> chemical pesticides<br />
203<br />
Immunotoxicity testing, direct food<br />
additives 202<br />
Immunotoxicology 196<br />
collaborative studies 197, 198<br />
screening tests 196<br />
Insulin-like growth factor 1 (IGF,) 265<br />
Interferon (IFN- ) 140–6<br />
Interleukin 3 (IL-3) 141,143
Interleukin 4 (IL-4) 140, 141, 143<br />
Interleukin 5 (IL-5) 141–6<br />
Interleukin 10 (IL-10) 141<br />
Interleukin 12 (IL-12) 142<br />
International Agency for Research on<br />
Cancer (IARC) 91, 93–7, 163<br />
International Programme on Chemical<br />
Safety (IPCS) 91<br />
Iron 246<br />
Isocyanates 151–6<br />
Isophorone diisocyanate (IPDI) 151<br />
Joint Expert Committee on Food<br />
Additives (JECFA) 208<br />
Kainic acid (KA) 242–7<br />
Labelling 296–9<br />
Lactate dehydrogenase (LDH) 117<br />
Late respiratory systemic syndrome<br />
(LRRS) 152–7<br />
Lead 244–8<br />
Leaving group 65–8<br />
Life cycle exposure 283<br />
Life span correction 47–49<br />
Light microscopic histopathology 123–6<br />
Light stabilisers 316–39<br />
benzotriazole-based 322–9<br />
Linearized multistage cancer model<br />
(LMS) 170<br />
Lipophilic compounds 206<br />
Liver enzyme induction 318–2<br />
benzotriazole-based light stabilisers<br />
327–31<br />
LOAEL 168, 367<br />
Local lymph node assay (LLNA) 196<br />
LOEL 43<br />
Low molecular weight (MW) organic<br />
chemicals 187<br />
Lung burden analysis 103–8, 121–5<br />
Lung digestion/biodurability studies<br />
125<br />
Lung dissection 118<br />
Lung fibre burden 98–1<br />
Lung tissue, fibre recovery from 118<br />
MAK-list 304<br />
INDEX 377<br />
Maleic anhydride (MA) 152<br />
Malformations 291<br />
Malignancy, critical mutations leading<br />
to 190<br />
Manganese 245, 246<br />
Man-made vitreous fibres (MMVFs)<br />
animal inhalation studies 95–106<br />
carcinogenic potential 91–115<br />
cell culture studies 94<br />
comparison <strong>of</strong> human exposures<br />
used in rat chronic inhalation studies<br />
108–13<br />
epidemiological studies 91, 93–7<br />
implantation studies 95<br />
potential biological effects 91<br />
previous inhalation studies 106–9<br />
toxicologic studies 94<br />
Maximum life potential (MLP) 47–54<br />
Maximum tolerated dose (MTD) 170<br />
Meehs Formula 47<br />
Mercaptans 21<br />
Mercapturic acids 21–4, 184–90, 187–3<br />
toxicokinetics <strong>of</strong> 23<br />
urinary excretion 15, 25<br />
Metabolic activation 60–4<br />
Metabolism and toxicity 206–12<br />
Methamphetamine 241<br />
Methylene chloride 65<br />
Methylene diphenyldiisocyanate (MDI)<br />
151<br />
Methylmercury 241, 245<br />
N-methyl-N-nitrosourea (MNU) 263<br />
Michaelis-Menten kinetics 173<br />
Mitogenic stiraulation (ConA.LPS) 199<br />
Model neurotoxins 241–52<br />
Monitoring<br />
environmental (EM) 19–1<br />
human exposure 188<br />
in occupational toxicology 19–1<br />
polycyclic aromatic hydrocarbon<br />
(PAH) exposure 158–6<br />
see also Biological effect monitoring<br />
(BEM);<br />
Biological monitoring (BM)<br />
Monoclonal antibodies (Mabs) 185–2,<br />
318<br />
MPP + 241<br />
MPTP 241
378 INDEX<br />
mRNA analysis 211<br />
Mutagenic potency 190–6<br />
Mutagenicity, surfactants 349–2<br />
N 7 -deoxyguanosine (N 7 -dG) 186<br />
NADPH-cytochrome P450 reductase<br />
15<br />
Naphthalene diisocyanate (NDI) 151<br />
β-naphth<strong>of</strong>lavone 266, 267<br />
2-naphthylamine 40–3, 301<br />
Neuropathy target esterase (NTE) 238<br />
Neurotoxicity, biomarkers <strong>of</strong> 238–4<br />
Neurotoxicity assessment 285<br />
Neurotoxicity testing 237–56<br />
Neurotypic proteins 238–3<br />
Nitroarenes 74, 77<br />
NK activity 199<br />
NK test 200<br />
NOAEL 168, 174, 367<br />
NOEL 43, 366, 367<br />
Nucleophilic centres 73<br />
Occupational asthma 148–9<br />
chemical agents causing 148–5<br />
incidence 148, 151<br />
initial diagnosis 149<br />
Occupational toxicology, monitoring in<br />
19–1<br />
OECD guidelines 421 282–7<br />
OECD guidelines 422 282–91<br />
OECD single generation study 283<br />
17β-oestradiol 212<br />
Organic solvents<br />
absorption <strong>of</strong> 3–7<br />
biological monitoring <strong>of</strong> 2–3<br />
dermal uptake <strong>of</strong> 5–7<br />
pulmonary uptake <strong>of</strong> 3–5<br />
Organophosphate-induced delayed<br />
neuropathy (OPIDN) 238<br />
Organophosphate pesticides 237–2<br />
Parkinson’s disease 241, 245<br />
PBPK/PBTK models<br />
1,3-butadiene 16<br />
2-butoxyethanol 166–4, 172–80<br />
development <strong>of</strong> 31–3<br />
in risk assessment 170–80<br />
PCA-DNA adducts 187<br />
PCNB 267<br />
Pentafluorophenyl isothiocyanate<br />
(PFPITC) 76<br />
Pentafluorophenyl thiohydantoine<br />
(PFPTH) 77<br />
Perchlorate-discharge test 260<br />
Peroxisome proliferation 223–40<br />
carcinogenicity 224–30<br />
in rodent liver 223–8<br />
mechanisms <strong>of</strong> 226–1<br />
risk assessment 229–5<br />
rodent 223–9<br />
species differences in response 227–3<br />
Pesticides 203, 202<br />
human exposure to 208<br />
immunotoxicity <strong>of</strong> 203<br />
organophosphate 237–2<br />
Pharmaceuticals 202<br />
Phencyclidine 48<br />
Phenobarbital 267, 273<br />
Phenobarbitone 322<br />
Phenolic antioxidants 316–25<br />
blood kinetics 317–1<br />
blood metabolites 317–1<br />
effects on serum thyrotropin and<br />
thyroid hormones 318–4<br />
liver enzyme induction 318–2<br />
model compound 317<br />
risk assessment 321–5<br />
Phthalate esters 223, 224<br />
Phthalic anhydride 152<br />
Physiologically-based pharmaco<br />
(toxico-) kinetic models. See PBPK/<br />
PBTK models<br />
Phytopharmaceuticals 208<br />
Pigments. See Colorants (dyes and<br />
pigments)<br />
Plaque assay (PFCA) 199, 200<br />
Plasticisers 223, 229–4<br />
Polychlorinated biphenyls (PCBs) 247–<br />
2, 267<br />
Polycyclic aromatic hydrocarbons<br />
(PAHs) 74–7, 78, 158, 186<br />
biomonitoring exposure 158–6<br />
Polyisocyanates 151–6<br />
Postlabelling 78–79, 81<br />
Post-natal manifestations 284
Post-radiolabelling technology 184<br />
Prenatal effects 284, 285, 289<br />
Production volume triggers 280, 281<br />
Propylthiouracil (PTU) 260<br />
Protein adducts 73, 74, 79, 81, 184–90<br />
Protein-binding 258<br />
Pulmonary cell proliferation 118, 124–7<br />
Pulmonary hyperreactivity to industrial<br />
pollutants 128–9<br />
guinea-pig model 131–6<br />
rat model 134–7<br />
Pulmonary sensitization, mechanisms <strong>of</strong><br />
137–51<br />
Pulmonary toxic effects<br />
chrysotile asbestos fibres 116–30<br />
para-aramid fibrils 116–30<br />
Pulmonary uptake <strong>of</strong> organic solvents<br />
3–5<br />
Pyrethroid insecticides 238<br />
Pyromellitic dianhydride (PMDA) 152<br />
Quinone-thioethers 63<br />
Quinones 63–6<br />
Rad-equivalence values 192<br />
Radioallergosorbent tests (RAST) 138<br />
Radioimmunoassays 77<br />
Reactive airways dysfunction syndrome<br />
(RADS) 128, 148, 151, 153<br />
Reactive chemicals, metabolism 60–71<br />
Reactive compounds<br />
adduct formation <strong>of</strong> 74<br />
determination in unknown mixtures<br />
83–6<br />
determination <strong>of</strong> 71–88<br />
interaction with cellular constituents<br />
71–5<br />
Reactive metabolites 71<br />
Refractory ceramic fibres (RCFs) 91,<br />
99–2<br />
industrial hygiene studies 105<br />
Repeated dose toxicity 281<br />
Reproductive toxicity 279–298<br />
detecting effects on males 293<br />
evaluation for 282<br />
interpretation/extrapolation <strong>of</strong> 296–<br />
9<br />
INDEX 379<br />
labelling 296–9<br />
manifestations <strong>of</strong> 291–4<br />
methods for detecting effects on 282<br />
overlap <strong>of</strong> 281<br />
Respiratory allergic hypersensitivity<br />
137<br />
Restricted test systems 283<br />
Rifampicin 322<br />
Risk assessment 162–9, 166–84<br />
and exposure pr<strong>of</strong>iles 362–68<br />
approaches to 167–6<br />
cancer 188–6<br />
guidelines for 361<br />
in vitro approach 207–13<br />
mathematical procedures 369–3<br />
non-threshold effects 368<br />
peroxisome proliferation 229–5<br />
phenolic antioxidants 321–5<br />
physiologically based<br />
pharmacokinetic (PBPK) models in.<br />
See PBPK/PBTK models<br />
textile chemicals 315<br />
threshold effects 366–71<br />
see also Safety assessment<br />
RNA probes 211<br />
Rock (stone) wool 91, 102–5<br />
industrial hygiene studies 108<br />
R-phrases 364<br />
Safety assessment 361–74<br />
benzotriazole-based light stabilisers<br />
331–7<br />
see also Risk assessment<br />
Salmonella typhimurium 66<br />
SHIELD system 149<br />
Slag wool 91, 102–5<br />
industrial hygiene studies 108<br />
Solid phase assays 77<br />
Sperm analysis 284<br />
Spermatogenesis 284, 286, 288<br />
Structure-activity databases 282<br />
Structure-activity relationships,<br />
surfactants 344<br />
Styrene 38–38<br />
Styrene oxide 38–38<br />
Superoxide dismutase (SOD) 241<br />
Surface markers 200
380 INDEX<br />
Surfactants 338–55<br />
acute toxicity 348–1<br />
biochemical properties 338–5<br />
carcinogenicity 350<br />
chronic toxicity 349<br />
embryotoxicity 350<br />
excretion 347–50<br />
interactions with enzymes 341–5<br />
interactions with membranes 339<br />
interactions with proteins 339–3<br />
intestinal absorption 347–50<br />
local effects 342–48<br />
metabolism 347–50<br />
mucous membrane compatibility<br />
345<br />
mutagenicity 349–2<br />
oral toxicity 348<br />
percutaneous absorptions 346–9<br />
sensitization 345–8<br />
skin compatibility 342–7<br />
structure/activity relationships 344<br />
systemic effects 348<br />
toxicokinetics 346–50<br />
Surveillance <strong>of</strong> Work-related and<br />
Occupational Respiratory Disease<br />
Project (SWORD) 148<br />
Synaptophysin 242<br />
Systemic bioavailability <strong>of</strong> colorants<br />
301<br />
Systemic toxicity assessment 285<br />
Target dose determination 188–4<br />
Temelastine 271 272, 273<br />
2-tert-buty1–4-methoxyphenol (BHA)<br />
316<br />
Testosterone 212<br />
Tetrachloroethane 7<br />
Tetrachloroethene 8, 7<br />
Tetrachlorophthalic anhydride (TCPA)<br />
152, 153<br />
Textile chemicals 308–18<br />
finishing plant 309–13<br />
handling and processing 312–18<br />
irritant properties 311, 312<br />
new developments regarding<br />
toxicology 309<br />
oral toxicity 310<br />
process <strong>of</strong>f-gas 315<br />
risk assessment 315<br />
temperature effects 311<br />
toxicological pr<strong>of</strong>ile 311<br />
toxicology assessment needs 315<br />
Thalidomide 291<br />
T helper (Th) cells 140–5<br />
Thioethers 21–3<br />
Threshold effects<br />
classification 364–9<br />
risk assessment 366–71<br />
Thyroglobulin (TBG) 258, 260<br />
Thyroid<br />
endocrine toxicology 254–80<br />
tumours <strong>of</strong> 256<br />
Thyroid binding pre-albumin (TBPA)<br />
258–2<br />
Thyroid follicles 260<br />
Thyroid follicular capability 262<br />
Thyroid follicular cell hyperplasia and<br />
neoplasia, pathobiology <strong>of</strong> 262–7<br />
Thyroid function<br />
control <strong>of</strong> 258<br />
perturbation <strong>of</strong> 256–60<br />
Thyroid hormones 258, 260, 318–4<br />
Thyroid lesions, pathobiology <strong>of</strong> 258<br />
Thyroid neoplasia 262, 321<br />
Thyroid stimulating hormone (TSH)<br />
258, 260, 262, 263, 265, 320, 321,<br />
322<br />
Thyroid toxicity 256<br />
via H-P-T-L axis 265–76<br />
Thyroid tumorigenesis 264<br />
Thyrotrophin releasing hormone (TRH)<br />
258, 262<br />
Thyrotropin 318–4<br />
Thyroxine 260, 269<br />
accumulation 271, 273<br />
clearance <strong>of</strong> 267–76<br />
T lymphocytes 142<br />
Toluene 241<br />
2,4- and 2,6-toluene diisocyanate (TDI)<br />
151<br />
Toxicity, and metabolism 206–12<br />
Toxicity tests, EC annex VII and VIII<br />
280<br />
Toxicodynamic interactions 20<br />
Toxicokinetic interactions 20
Toxicokinetic parameters 16<br />
Toxicokinetics, principles <strong>of</strong> 15<br />
Transition metals 246<br />
Trichloroacetic acid (TCA) 28, 29<br />
Trichloroethane 7, 8<br />
Trichloroethylene (TRI) 26–31<br />
effects related to 26–31<br />
hepatocarcinogenicity induced by<br />
28<br />
oxidative metabolism 28, 29<br />
2,5,6-trichloro-3-glutathion-S-yl)<br />
hydroquinone 64<br />
Triethyl lead 241<br />
tri-(2-ethylhexyl)trimellitate) 223<br />
Trimellitic anhydride (TMA) 152, 153<br />
Trimethyltin (TMT) 241–6<br />
Tumour incidence 189–5<br />
Two generation studies 291–5<br />
alternatives to 295–8<br />
UDP-glucuronosyltransferase 321, 327,<br />
329<br />
UDP-glucuronyltransferase (UDP-GT)<br />
213, 260, 267, 271, 274<br />
University <strong>of</strong> Pittsburgh study 91, 92<br />
Urinary excretion<br />
mercapturic acids 15, 25<br />
xenobiotics 16–18<br />
US Environmental Protection Agency<br />
91<br />
US-NTP study 200<br />
Veterinary medicinal chemicals 202<br />
Vinyl chloride 36–9<br />
Western blotting 211<br />
World Health Organization (WHO) 91<br />
Xenobiotic-induced endocrine<br />
dysfunction 254<br />
Xenobiotics 11<br />
assessment <strong>of</strong> long-term toxicity<br />
207<br />
biological effects 14<br />
disposition <strong>of</strong> 12–15<br />
overall exposure to 20<br />
urinary excretion <strong>of</strong> 16–18<br />
INDEX 381