Literature review: species translocations as a tool for biodiversity ...

Scottish Natural Heritage
Commissioned Report No. 440
Literature review: species translocations as
a tool for biodiversity conservation during
climate change

COMMISSIONED REPORT
Commissioned Report No. 440
Literature review: species translocations
as a tool for biodiversity conservation
during climate change
For further information on this report please contact:
Dr David Genney
Scottish Natural Heritage
Great Glen House
INVERNESS
IV3 8NW
Telephone: 01463-725 253
E-mail: david.genney@snh.gov.uk
This report should be quoted as:
Brooker, R., Britton, A., Gimona, A., Lennon, J. & Littlewood, N. (2011). Literature
review: species translocations as a tool for biodiversity conservation during climate
change. Scottish Natural Heritage Commissioned Report No.440.
This report, or any part of it, should not be reproduced without the permission of Scottish Natural Heritage.
This permission will not be withheld unreasonably. The views expressed by the author(s) of this report
should not be taken as the views and policies of Scottish Natural Heritage.
© Scottish Natural Heritage/MLURI 2011

COMMISSIONED REPORT
Summary
Literature review: species translocations as a tool for biodiversity
conservation during climate change
Commissioned Report No. 440 (iBids n o 7966)
Lead author: Dr Rob Brooker, The Macaulay Institute, Aberdeen
Year of publication: 2011
Background
Climate change is a major threat to global biodiversity. Changes in the location of suitable
climatic conditions may drive species range-shifting, as species attempt to track suitable
climate through space. However, land use change and associated habitat fragmentation, as
well as inherent characteristics of the species (e.g. rate of reproduction and dispersal ability),
may place serious limitations on the ability of species to track suitable climate. Some species
will be unable to reach regions with suitable future climatic conditions, leading to the
proposed use of species translocations as a tool to overcome this stranding effect: species
at risk would be moved to those areas expected to be suitable for their growth. This
procedure is termed, here, Assisted Colonisation (AC). There is considerable current debate
concerning the application of AC as a conservation tool for dealing with the threat of climate
change.
The issues associated with the use of such translocations are directly relevant to species
conservation in Scotland. The current SNH-Macaulay partnership project, of which this
literature review is a part, aims to consider the application of species translocations as a
conservation tool for the establishment or protection of populations in northerly and/or
montane environments in Scotland. This literature review provides information to help focus
later stages of the project.
Main findings
Although there is no shortage of possible targets for research projects exploring the
application of AC, we would suggest that the following are priority areas for action:
Developing assessment processes to attempt to identify candidates for AC.
Investigations of dispersal abilities and genetics.
Practical trials of transplant methodologies for particular species groups.
Better understanding of what constitutes “last resort”.
Exploring the best locations from which to source material for AC relative to current and
future climatic conditions.
Developing predictive techniques for selecting sites that will be suitable for future survival.
For further information on this project contact:
Dr David Genney, Scottish Natural Heritage, Great Glen House, INVERNESS, IV3 8NW, Tel: 01463 725 253
For further information on the SNH Research & Technical Support Programme contact:
DSU (Policy & Advice Directorate), Scottish Natural Heritage, Great Glen House, Inverness, IV3 8NW.
Tel: 01463 725000 or pads@snh.gov.uk
ii

Table of Contents
EXECUTIVE SUMMARY
Page
iv
1 INTRODUCTION 1
1.1 Background 1
1.2 Aims of the literature review 1
1.3 Definitions 2
1.4 A summary of current debate concerning assisted colonisation 2
2 SPECIES GROUP NO. 1 - LICHENS & BRYOPHYTES 6
2.1 Threats to lichens & bryophytes in Scotland 6
2.2 Dispersal & establishment 7
2.3 Selecting target species 8
2.4 AC as a tool for conservation 8
2.5 Summary & suggestions for experimental studies 11
3 SPECIES GROUP NO. 2 – VASCULAR PLANTS 13
3.1 Threats to vascular plants in Scotland 13
3.2 Possible risks from assisted colonisation 16
3.3 Factors determining the practicality and success of assisted colonisation 17
3.4 Summary & suggestions for Experimental Studies 20
4 SPECIES GROUP NO. 3 – TERRESTRIAL INVERTEBRATES 22
4.1 Threats to terrestrial invertebrates in Scotland 22
4.2 Possible risks from assisted colonisation 23
4.3 Factors determining the success of assisted colonisation 24
4.4 Practicalities of translocating terrestrial invertebrates 26
4.5 Summary & suggestions for experimental studies 26
5 SPECIES DISTRIBUTION MODELS: PREDICTING SITES FOR ASSISTED
COLONISATION 28
5.1 Introduction 28
5.2 Species-habitat models 29
5.3 Climate envelope models 29
5.4 Landscape scale and local scale models 32
5.5 Mechanistic process-based models 33
5.6 Application of modeling approaches to assisted colonisation 35
6 UNOCCUPIED HABITAT – AN ALTERNATIVE TO AC? 38
6.1 Are there empty niches? 38
6.2 What is a species’ range? 39
7 UNDERSTANDING THE AC DEBATE, THE DECISION MAKING PROCESS, AND
THE POLICY CONTEXT 41
7.1 Understanding the AC debate and alternative decision-making approaches 41
7.2 Conservation policy and guidelines relevant to assisted colonisation 42
8 SUMMARY & SYNTHESIS 45
8.1 Areas for priority action 46
9 REFERENCES 47
iii

Acknowledgements
The authors would like to thank the following for extremely helpful comments on a previous
version of this document: Mairi Cole, SNH; Sarah Dalrymple, University of Aberdeen; David
Genney, SNH; Pete Hollingsworth, RBGE; Chris Wilcock, University of Aberdeen; Mark
Young, University of Aberdeen.
iv

Executive Summary
Background
Climate change is a major threat to global biodiversity (Hannah et al. 2002, Hoegh-Guldberg
et al. 2008a, b). Changes in the location of suitable climatic conditions will drive species
range-shifting, but various factors (e.g. land use change, species’ characteristics) will place
limitations on the ability of species to track suitable climate. This has led to proposals for the
use of a particular type of species translocations for conservation during climate change:
“stranded” species would be moved to areas expected to be suitable for their growth.
However, there is considerable debate concerning the application of such assisted
colonisation (AC).
Questions surrounding the use of AC are directly relevant to species conservation in
Scotland. This literature review provides information on the current state of understanding
concerning the use of AC, and has been used to focus later research stages of the
associated Macaulay Institute-SNH partnership project.
Debate concerning the application of AC
Debate concerning the application of AC is driven by a number of factors, including:
Selection of target species and individuals, and impacts on donor populations.
Determination of the point of “last resort”.
Impacts on the recipient community including species invasions and pathogens.
Detraction of effort from other conservation actions.
We considered these factors in the context of the application of AC in Scotland with respect
to three species groups, lichens and bryophytes, vascular plants, and terrestrial
invertebrates.
Lichens & Bryophytes
Climate is a major factor controlling the distribution of cryptogams (lichens and bryophytes)
in the UK. Climate change will interact with other factors (e.g. habitat loss and fragmentation,
atmospheric deposition of pollutants) to affect the availability of suitable cryptogam habitats.
The difficulty for conservation efforts will be to identify those rare species which are limited
by dispersal rather than establishment or habitat availability. Autecological information for
both common and rare cryptogam species is generally scarce, making such predictions
difficult. Lichens or bryophytes very rarely become invasive, and it is unlikely that potential
AC targets would share the attributes of the known invasive cryptogams.
Transplantation experiments demonstrate that obtaining a successful physical bond between
transplanted material and the recipient substrate, detailed knowledge of the species’
autecology, and propagule type, all influence success. There is currently a mismatch
between the habitats for which we have knowledge of species’ dispersal and those which
appear likely to be most impacted by climate change. Studies of arctic-alpine species would
be very useful, as would practical trials of transplant methodologies.
Vascular plants
Determining which vascular plants are candidates for AC is complex. Arctic-alpine species
are highly threatened by climate change, but suitable climate space may not be available for
them in future in Scotland. At lower altitude, impacts of land use change and habitat
fragmentation make it hard to prove that climate change is driving species declines. Data on
dispersal of vascular plants and evidence of in situ adaptation are lacking. In Scotland, the
risk of invasion of recipient communities by translocated plant species might be considered
low. The possible occurrence of hybridisation following translocation is relatively predictable.
v

Propagule type and source population location, selection of recipient sites, and the recreation
of essential interactions are important factors in determining success of vascular
plant translocations, but even for this well-studied group there is a limited knowledge base.
Recommendations for further research include: assessing which species are likely to be
candidates for AC; better population-level monitoring; better understanding of what
constitutes “last resort”; exploring the best locations for donor and recipient populations or
communities.
Terrestrial invertebrates
As for other groups, defining which terrestrial invertebrates may be threatened by climate
change and aided by AC is difficult. Climate-driven changes to insect distributions may be
hard to detect as accurate data are lacking. Greater threats clearly exist for habitat
specialists in isolated patches, while local adaptation might also limit dispersal. In such
cases AC may be the only conservation action available, but recipient sites need to be
prepared years or decades in advance. Phytophagous invertebrates have clear potential to
become invasive. A low-risk strategy might be introductions into areas formerly occupied and
which are now suitable under a changing climate, or to sites close to a species’ origin.
The factors determining the success of insect introductions vary considerably on a case-bycase
basis, although information on the long-term success of such translocations is scarce.
Habitat suitability is crucial, and necessitates knowledge of a species’ autoecology. Insect
(re)-introductions can also be hampered by lack of fitness of the stock used. Key
uncertainties are: processes for identifying suitable AC candidates; the relationship of
introduced target species to parasites; environmental factors currently limiting focal species’
range; autecology of target species.
Species distribution models: predicting sites for assisted colonisation
To predict climatically suitable current and future locations we might adapt two main groups
of models already utilised to explore the impacts of climate change on biodiversity. Empirical
models include species-habitat models and climate envelope models, which differ only in
spatial scale. Climate-envelope models (CEMs) utilise relationships between present climatic
variables and distribution data but, although widely used, are limited in what they can tell us.
Their various assumptions are unrealistic (e.g. equilibrium between species distributions and
climate), and statistical problems mean that they should be interpreted, at best, as indicating
potential future distributions.
Mechanistic process-based models incorporate greater realism by including, for example,
biological processes and biotic interactions, and can be combined with CEM models to
predict realised rather than potential future distributions. However, getting sufficient
information to build realistic models is an expensive and time-consuming task. Pragmatic
steps to improve the situation include development of better species distribution modeling
approaches (e.g. CEMs) linked to common garden experiments, or a triage process (based
on a range of criteria) for selecting suitable species for AC.
Unoccupied habitat and the definition of range
An alternative to AC might be the translocation of individuals within the species’ current
range to sites that are suitable but currently unutilized. The first key problem is determining
whether that location will remain suitable in the future. The second key problem is
determining whether the recipient site is inside of a species’ range, as clear definitions of
range are commonly lacking. An important distinction is that between the extent of
occurrence (EoO) and the area of occupancy (AoO) - the AoO may not contain areas of
suitable but unoccupied habitat, despite the EoO doing so. The scale at which range is
plotted can also be influential. There is a clear need to agree how ranges are defined in
order to appropriately classify any translocation activity.
vi

Understanding the AC debate, the decision making process, and the policy context
The AC debate is complex and nuanced. A number of frameworks have been proposed for
working through the decision making process. Several studies propose linear decision trees,
but these have been criticised as simplistic, with simultaneous consideration of costs and
benefits using multidimensional evaluation frameworks being proposed as an alternative.
Irrespective, it would seem sensible for each decision concerning AC to be made on a caseby-case
basis but following some form of standardised guidelines.
Conservation guidelines covering species translocations provide a number of stages that
should be worked through when planning a re-introduction which, with the exception of
limiting release only to sites within the historic range, are directly applicable to AC. Additional
injunctions are recommended on the use of AC, particularly that it “should be undertaken
only as a last resort”. The need or desire for new legislation to regulate AC appears driven
by a fear of unregulated AC. However, such actions might be averted if there was a greater
public understanding of the risks to in situ species and habitats.
Summary & synthesis
The problems of applying AC in Scotland are not unique. Species selection is particularly
complex, but might be based on a combination of both assessed threat from climate change,
and characteristics that make the application of AC acceptable and likely to succeed. These
latter include AC being the “last resort”, predicted low impact on donor populations, selection
of sufficiently diverse and suitable genetic stock and of appropriate recipient sites, the
recreation of necessary biotic interactions, and a low impact on recipient communities.
However, although such characteristics are easy to list, their assessment is complex, the
central over-arching problem being the lack of data for many potential candidate species.
Although there is no shortage of research targets, priority areas for research action include:
developing assessment processes to identify candidates for AC; investigations of dispersal
abilities and genetics; practical trials of transplant methodologies; better understanding of
what constitutes “last resort”; exploring the sourcing of material for AC; developing predictive
techniques for selecting sites that will be suitable for future survival.
vii

1 INTRODUCTION
1.1 Background
Climate change is clearly one of the major threats to global biodiversity (Hannah et al. 2002,
Hoegh-Guldberg et al. 2008a). Changes in climatic conditions alter the suitability of
environments for species, and under future climate scenarios the occurrence of suitable
climatic conditions may or may not overlap with a species’ current range. Even if some parts
of a species’ current range and the location of future suitable climatic conditions overlap, the
degree of overlap is likely to be highly variable between species.
Changes in the location of suitable climatic conditions may drive species range-shifting, as
species attempt to track suitable climate through space. However, other environmental
change drivers such as land use change and associated habitat fragmentation, as well as
inherent characteristics of the species such as their rate of reproduction and dispersal ability,
may place substantial limitations on the ability of species to track suitable climate.
Consequently some species may become “stranded”, i.e. unable to reach those regions
where suitable future climatic conditions exist. This has led to the proposed use of species
translocations as a tool by which conservationists can overcome this stranding effect. Put
most simply, species at risk would be moved from their current locations to those areas
expected to be suitable for their growth under future climate change scenarios. However,
although an apparently simple concept, there is considerable current debate concerning the
application of translocation as a conservation tool for dealing with the threat of climate
change.
The issues associated with the use of such translocations are directly relevant to species
conservation in Scotland. There is concern that climate change represents a considerable
threat to a wide range of Scottish species (Brooker et al. 2004). In addition, habitat
fragmentation and the barriers that this places on dispersal processes have also been
identified as a key conservation issue in Scotland (SNH 2009).
1.2 Aims of the literature review
The current SNH-Macaulay partnership project, of which this literature review is a part, aims
to consider the application of species translocations as a conservation tool for the
establishment or protection of populations in northerly and/or montane environments.
Because of the complexity of the issues associated with this conservation technique, and its
relatively uncertain status, the literature review provides useful information to help focus later
field-based experimental stages of the project. First, the review discusses some of the
current generic debate concerning translocations and climate change, and then considers
the application of this technique to three species groups: lichens and bryophytes, vascular
plants, and invertebrates. It is likely that the issues associated with translocation differ
between groups, hence the benefit from breaking our assessment into these subsections.
These groups have been chosen because 1) the research team assembled for this project
have experience of working with them, 2) they are likely to be more readily amenable to
field-based experimental trials than, for example, vertebrate animals.
We also consider the application of various ecological modeling techniques to the issue of
species translocation and climate change (Section 5). Computer-based modeling has been
widely used in assessments of the impacts of climate change on biodiversity, and attempts
to understand possible biodiversity responses. However, considerable debate also rages in
this arena concerning the applicability of different modeling approaches and the simplifying
assumptions involved in the modeling of complex ecological systems. A separate section of
1

the literature review therefore examines these various modeling approaches and discusses
their applicability to the question of species translocations in a climate change context.
Finally we attempt to synthesize the information gathered from the literature review,
providing a summation of commonalities or clear differences between the possible use of
translocation for the three species groups, the applicability of ecological modeling to them,
and key research questions that might be addressed by any experimental study.
1.3 Definitions
The terminology associated with species translocations is not straightforward (Armstrong &
Seddon 2008). The IUCN position statement on translocation (IUCN 1987) sets out three
main categories of translocation:
1. Restocking (often called reinforcement): translocations used to bolster existing
populations of plants or animals within their current range, for example to deal with
inbreeding depression or dispersal-limitation between populations.
2. Re-introduction: intentional movement of an organism into part of its known range from
which it has disappeared or been extirpated during historic times.
3. Introduction: movement of an organism outside of its historically known native range.
Most likely because they were written prior to current discussions of translocation within the
context of climate change, the 1987 IUCN categorisation does not mention the wide range of
names currently being used to describe activities that fall within this third category. In
particular, introductions which place a species outside of its current range, and which are
undertaken with the aim of either placing a species within future suitable climate space or
enhancing its ability to reach such space (e.g. by overcoming dispersal barriers and
enhancing climate tracking), are called assisted migration, assisted colonisation (or
colonization), artificial dispersal, artificial introduction, managed relocation, conservation
introductions, or even planned invasions (Ricciardi & Simberloff 2009a). The choice of this
more recent terminology appears to reflect in some cases the support given to this
conservation technique by different authors, or the particular species group being discussed.
Hunter (2007) recommends using the term assisted colonisation as “migration” has particular
connotations for zoologists. Throughout this literature review we will use the term assisted
colonisation, hereafter referred to as AC.
1.4 A summary of current debate concerning assisted colonisation
Recently, the application of AC has been intensely debated. However, this is not because
the application of such translocations is new. Examples can be found throughout the
conservation literature of species that have been moved to regions outside of their known
range (see, for example, Maunder 1992). Often, this has been because the translocated
species has a particularly restricted local distribution (e.g. endemics) and their known
populations are threatened by some specific (often anthropogenic) factor such as
urbanisation or the impact of invasive species. In such cases, where the original natural
habitats of the species have been destroyed, the translocation is described, perhaps
euphemistically, as a “benign introduction” (Seddon et al. 2009). In some other cases,
however, translocations have occurred because the translocated species has the potential to
perform some useful ecosystem function, for example herbivore-based scrub control
(Hayward 2009).
2

The recent, and at times heated, debate has probably arisen because of two separate
factors. Firstly, there is increasing recognition of the likely inability of many species to be
able to adapt in situ to climate change, particularly if population genetic diversity has been
eroded (Skelly et al. 2007). Similarly, species may be unable to undertake range shifting to
track suitable climatic conditions through space, for example through poleward or upward
movement. Although species’ ranges are often climate-influenced and have tracked previous
climatic fluctuations, the current rate of climate change, coupled with the massive
fragmentation of natural systems in many locations, leads to a sizable mismatch between
the necessary rate of range shifting and the actual capacity of species to move (Travis
2003). As a consequence, some conservation scientists are starting to consider more
interventionist techniques in order to protect species during climate change, including largescale
seed or gene banking (e.g. Vitt et al. 2009) or AC.
Secondly, at the same time as the potential scale of the detrimental impacts of climate
change on biodiversity were becoming recognised, considerable attention was also being
given to the impacts of invasive species. After land use change, some analyses (e.g. Sala
2000) consider invasive species to be the second most pressing threat to global biodiversity.
Many studies have examined whether it is possible to predict which species become
invasive, but generally conclude that this is not straightforward, not least because
mechanisms leading to invasion include “novel weapons” (Callaway & Ridenour 2004), or
the rapid evolution of species in response to their new environmental conditions (e.g. Davis
2005). Consequently, suggestions that AC might be used as a species conservation tool
have been met with alarm by researchers, conservation practitioners, NGOs and policy
makers concerned about the unpredictable nature of the impacts of translocated species on
the recipient communities.
However, these are not the only concerns about the possible use of AC. Notably some of the
concerns raised are about ecological processes, others about the strategic allocation of
limited conservation resources and the impact of AC on public perception of biodiversity
conservation. Below we summarise some of the major points of contention concerning AC:
1.4.1 Selection of target species and individuals, and impacts on the donor population
Discussions concerning the application of AC generally agree that it should be used as a
“last resort” when other conservation strategies have failed, or are clearly likely to.
Application of AC would require risk assessments prior to the translocation itself.
Translocations to new sites of species with highly restricted distributions have been
conducted under such circumstances, for example when development threatens native
habitat. In such cases the fact that the point of “last resort” has been reached is relatively
clear – there is no doubt that the existing sites will be lost. With respect to AC, however, this
is a highly contentious decision. As will be discussed in Section 5, predicting the movement
(or not) of species in response to climate change is extremely complex, and becomes
increasingly difficult as species distributions (and hence data for modeling) become more
restricted and fragmented. Furthermore, for many of the rare species which are likely to be
the initial focus of any AC assessments, we lack a fundamental understanding of why they
are rare, and so it becomes difficult to indisputably conclude that a single driver, climate
change, will lead inevitably to their extinction (Ricciardi & Simberloff 2009a).
Reviews of re-introductions suggest that a major factor contributing to the success of
translocated populations is that the propagules come from donor sites with relatively large
populations (Griffith et al. 1989). There may be a number of mechanisms underlying this
effect, including genetic processes such as the risk of inbreeding depression or restriction of
genetic diversity (and hence capacity to adapt to the new site) within small donor
populations. Furthermore, larger and probably more robust populations will be better able to
donate the propagules that will be translocated. Successful re-introductions frequently
necessitate repeated translocations, creating a need for repeated extractions from the donor
3

population (Mueller & Hellmann 2008). One way to avoid this may be ex situ propagation of
material for translocation, a technique which is commonly applied during re-introduction
programmes. However, ex situ propagation has its own difficulties (see Section 3.3), and
cannot regenerate lost genetic diversity. Therefore, waiting until AC becomes a technique of
last resort may substantially reduce the probability that it will be successful, as well as
having a greater impact on the donor population. In addition it limits our ability to assess the
impact on donor populations. For example, if only a single population is left its extinction
following donation would be hard to attribute to the act of donation (Griffith et al. 1989)
because there would be no untouched (non-donating) control population against which to
compare it.
A further issue in undertaking AC is the selection of individuals - rather than species - for
translocation. Some AC discussions have focused on the locations within a species current
range from which individuals for translocation should be selected (e.g. Hoegh-Guldberg et al.
2008a). These discussions have considered whether it is better to select from across a
species’ range, increasing the genetic diversity in the founder population, and hence its
adaptability, or whether it is better to select individuals from those populations most tolerant
of high temperatures (e.g. the southern limits of a species range for northern hemisphere
species) and hence ensure that the translocated population is most able to cope with high
temperatures. However, outbreeding depression might result if individuals from a number of
locally-adapted populations are brought together at a single introduction site.
1.4.2 Impacts on the recipient community
As discussed, one clear risk from AC is that translocated species become invasive.
However, uncertainty exists concerning the assessment of the risk that is posed by
invasives. For example, Gurevitch and Padilla (2004a) suggested that the impact of nonnative
invasives on native biodiversity may be overestimated: many species driven to
extinction are influenced by multiple detrimental drivers, and the presence of non-native
species may simply be correlated with, rather than driving, species extinctions. As Gurevitch
and Padilla (2004a) state “Exotic species might be a primary cause for decline, a
contributing factor for a species already in serious trouble, the final nail in the coffin or
merely the bouquet at the funeral". Their conclusions were criticised for underplaying the
possible impact of invasive species (Ricciardi 2004), although Gurevitch & Padilla (2004b)
defend their original assessment as being valid for the data available. Other assessments of
invasion risk, for example a US-wide analysis of invasive species by Mueller & Hellmann
(2008), have found that intracontinental translocation appears to be less risky, but that the
risk also differs between species groups (with, for example, relatively less risk from vascular
plant introductions compared to introductions of freshwater vertebrates). Analyses of
phylogenetic distance between native communities and invasive species in California
(Strauss et al. 2006) has shown that greater evolutionary divergence between the introduced
species and the recipient community increases the risk of an introduced species becoming
invasive.
However, although some factors such as reduced evolutionary or geographic distance may
reduce the risk of invasive impacts by non-native species, they do not absolutely eliminate it.
Introduced species still become invasive in some circumstances where the risk might be
considered low, for example following the intracontinental movement of black locust -
Robinia pseudoacacia – in the US. Such events remain to a large extent unpredictable
because of the context-specific and unique nature of some of the processes involved. As
Ricciardi & Simberloff (2009b) put it “contingency is the greatest impediment to prediction”.
An additional risk to the recipient community is the transport of pathogens along with the
introduced stock (Hodder & Bullock 1997). This might be particularly true for organisms held
or bred in captivity, as they will be exposed to a potentially greater range of pathogens than
in their native habitat. Mechanisms can be put in place to reduce this risk, for example
4

thoroughly cleaning roots of cultivated plants and/or treating with appropriate pesticides prior
to transplanting (see Section 3) or quarantining of animals (see Section 4), but it may be
very difficult to negate the risk of pathogen transfer. Some researchers have also suggested
that the translocation of associated pest species may be critical in maintaining ecological
balance in the recipient system – the release of species from the enemies encountered in
their native ranges is considered one of the major drivers of introduced species becoming
invasive (Callaway & Maron 2006).
1.4.3 Other conservation actions may represent a better investment of resources
Numerous techniques can be used for promoting species survival, of which AC is only one.
Choosing between these different management options is effectively a balance of the costs
and benefits of the different techniques, and such assessments can be extremely subjective
and influenced by the inherent prejudices of the assessor.
Some researchers argue that AC will be a relatively expensive technique - given the
necessity of detailed studies of the species, as well as donor and recipient systems - and
that the investment of funds in AC for a single species would be better spent on habitat
improvement to help save multiple alternative threatened species. Certainly, for any given
species, prior to AC becoming the preferred conservation option substantial effort should
have been put into conservation in situ unless it is certain that such effort would be wasted
(although, as discussed, under what circumstances would we have such certainty). The
hope is that this would greatly reduce the number of species for which AC will become
necessary. For example, Hannah et al. 2002 state that "Artificial translocation [i.e. AC],
previously proposed as the primary conservation response capable of keeping pace with
human-induced climate change… can be minimized with the careful design of dynamic
conservation systems that operate on a landscape scale". It has to be asked, though, what
these dynamic conservation systems actually are. For example, the impacts of habitat
fragmentation are well known (e.g. Wilcock & Neiland 2002, Travis 2003), but the proposed
use of increasing landscape connectivity as a conservation tool to deal with fragmentation
masks our lack of understanding as to how we might actually achieve this connectivity for
the range of species affected (Hodgson et al. 2009).
The costs of AC, as well as our ability to find workable alternatives, are also a point of
debate. Some of these costs can clearly be assessed in monetary terms. However, during
their assessment it should be remembered that the monetary costs associated with acquiring
detailed species-specific knowledge prior to AC will also be incurred as part of “traditional” in
situ or ex situ conservation, and so it is the additional costs which are specific to AC (e.g.
from surveying recipient sites and moving propagules) which should be the point of debate.
Furthermore, non-monetary valuations and costs are more difficult to define. For example,
Davidson and Simkanin (2008) argue that, during AC, it is often implicitly assumed that the
“value” of the translocated species is much greater than that of the recipient community.
But, as has been pointed out during recent discussion of valuation of ecosystem services,
non-monetary valuation of the “worth” of different elements of biodiversity can be extremely
difficult (Layke 2009). Another hypothetical cost whose genuine impact is hard to judge is the
possible gradual erosion of investment in alternative in situ conservation techniques. Even if
AC starts off as a tool of last resort, it might become more acceptable over time (particularly
if it can be shown to be successful), hence eroding support for in situ conservation efforts
(Ricciardi & Simberloff 2009a) and possibly the legislation that promotes them. Furthermore,
AC can be seen as a distraction, as it fails to address the root cause of species decline
(Fazey & Fischer 2009).
5

2 SPECIES GROUP NO. 1 - LICHENS & BRYOPHYTES
2.1 Threats to lichens & bryophytes in Scotland
Climate is a major factor controlling the distribution of cryptogams (lichens and bryophytes)
in the UK due to the high spatial variability in key variables such as rainfall and temperature.
As a result the flora contains species groups associated with many different climatic zones
e.g. temperate, boreal and oceanic, some of which, e.g. arctic-alpine cryptograms, are at the
edge of their range in the UK. Human-induced rapid climate change has the potential to
have major impacts on the distribution of these species groups. Climate envelope modeling
of the UK lichen flora (Ellis et al. 2007a) suggests that northern-montane and northernboreal
groups are particularly under threat from loss of climate space, while the effects on
other key groups such as the internationally important oceanic flora are less clear and could
be positive or negative (Ellis et al. 2007a, b). Climate change does not, however, act in
isolation and will interact with other factors affecting the availability of suitable cryptogam
habitats including direct habitat loss and fragmentation through, for example, land use or
land management changes and more diffuse impacts such as aerial deposition of pollutants
including nitrogen (N) and sulphur (S). All three drivers will act together to control the
distribution and abundance of cryptogam species, as has been shown for lichens on juniper
scrub (Ellis & Coppins 2009).
Habitat loss and fragmentation can, for some species, be key factors affecting the viability of
populations and their ability to respond to climate change. Many cryptogam species are
thought to be reliant on long-term habitat continuity for maintenance of their populations
(Coppins & Coppins 2002), particularly species which are dispersal-limited or those with very
narrow niches and specialist habitat requirements. Much work regarding the effects of
habitat fragmentation has been undertaken in woodland habitats, where it has been shown
that loss of habitat area may have long-term consequences, with species richness declines
continuing over a period of 10s-100s of years as a result of a so called ‘extinction debt’ (Ellis
& Coppins 2007). Fragmentation of habitats also has the potential to restrict gene flow
between populations leading to genetic isolation and reduced fitness, and may reduce the
ability of species to move through the landscape tracking climate changes.
Aerial deposition of pollutant nitrogen and sulphur is also a major factor causing ongoing
changes in vegetation composition within the UK (NEGTAP 2001). Pollutant deposition is a
particular problem for poikilohydric bryophytes and lichens which take up both water and
nutrients over their whole surface with much less control than higher plants. This renders
them sensitive to the effects of excess nutrient deposition which can have direct toxic effects
at high loads or concentrations (Britton & Fisher 2007, Pearce & van der Wal 2008, Britton &
Fisher 2010). Nitrogen deposition may also have indirect effects through stimulation of the
growth of higher plants and a resulting increase in competition for light which negatively
impacts low growing cryptogams (Cornelissen et al. 2001). Even though Scotland is
perceived to be a clean air region of the UK, many habitats receive deposition loads in
excess of the Critical Loads above which damage is likely to occur. Alpine habitats and
woodlands are especially vulnerable to the effects of nitrogen deposition as a result of high
rainfall and prolonged cloud cover in the former, and scavenging of pollutants by the tree
canopy in the latter (NEGTAP 2001). Thus nitrogen deposition will have large scale impacts
on habitat suitability for cryptogams, further decreasing the available space in which to find
new areas of suitable habitat in the face of climate change.
When planning responses to climate change for cryptogam conservation, all three of these
drivers (N deposition, habitat fragmentation and climate change) will need to be taken into
account since all will have large impacts on the amount and distribution of suitable habitats
within the landscape. Understanding species’ abilities to disperse around the habitat matrix
6

created by these three drivers will be critical for identifying those species most at risk of
(local) extinction as a result of the changes.
2.2 Dispersal & establishment
Bryophytes and lichens tend to show wider geographic ranges than other terrestrial plants,
with many species having extremely wide geographic ranges within climatic regions. Even
when suitable environments have large disjuncts in their distribution many species
distributions follow the same patterns (Schofield 1992). Such distributions, coupled with the
fact that many cryptogams are characteristic of temporally dynamic habitats, suggests a high
dispersal capacity. There has been some recent research which has investigated the
dispersal ability and metapopulation dynamics of cryptogams in fragmented landscapes and
which may be relevant for predicting species’ ability to respond to climate change. Much of
this research has focused on the dynamics of epiphytic species within fragmented forest
landscapes and on the dispersal characteristics of rare vs. common species, but general
principles may be applicable to a broader range of habitats.
Both bryophytes and lichens have multiple methods of dispersal, including a variety of
sexual and asexual methods. Bryophytes may reproduce sexually by spores or asexually by
means of fragmentation (deciduous shoot tips, leaves etc) or the production of bulbils, tubers
or gemmae. Lichens also reproduce by spores or by vegetative means including soredia,
isidia or larger thallus fragments. For lichens there is the added complication that only the
fungal partner produces the spores, and although a small number of algal cells may be
transported along with a spore, this mode of reproduction usually requires suitable algal cells
to be present at the destination. Soredia, isidia and thallus fragments contain cells of both
partners and are capable of forming a new lichen in any suitable habitat. The main contrast
between spores and other reproductive structures for both groups is their size; while
vegetative fragments are relatively large and are thought to be transported over relatively
short distances (in the order of 100s of metres; Werth et al. 2006), their size also means that
establishment rates will be higher and they may thus be an effective means for local spread.
Spores by contrast are generally very small and are supposedly easily transported. Since
they are produced close to the ground due to the small stature of most cryptogams, the
majority of spores fall close to the parent plant (During 1992), but those that escape the
boundary layer into free air may be transported over very large distances, as seen in spore
trapping studies in Antarctica which have detected moss spores arriving from Chile (Marshall
& Convey 1997). Despite their potential for ensuring efficient long distance dispersal, many
species do not, or only rarely, produce spores and spore production often occurs much later
in the life cycle than production of vegetative propagules which commonly starts soon after
establishment (Löbel et al. 2009). It has been suggested that species at the edge of their
range occupying suboptimal habitats are also less likely to produce spores (Cleavitt 2005);
this may be due to an energetic cost, since spore production is sometimes associated with
reduced growth (Löbel & Rydin 2009).
Several studies have looked at the distribution of different modes of reproduction between
common and rare species in an effort to generalise about the degree to which rare species
are dispersal or habitat-limited. In the UK bryoflora, Longton (1992) found that a higher
proportion of rare species did not produce spores (55%) than in the total flora (19%). A
positive correlation has been seen between spore production and range size in some
regions of the UK (Callaghan & Ashton 2008) and in Scandinavian epiphytes declining spore
size was associated with increasing scales of spatial aggregation (Löbel et al. 2009) which
might suggest better dispersal of small-spored species. However, spore production and size
is not always seen to be a good predictor of range size or habitat occupancy (Johansson &
Ehrlén 2003, Pharo & Zartman 2007, Leger & Forister 2009) so it may be difficult to
generalise dispersal ability from reproductive characters alone.
7

Recent molecular techniques, by measuring gene flow, have allowed more direct testing of
the ability of species to disperse through fragmented landscapes. These techniques have
mainly been applied to epiphytic species; although these species are thought to require
continuity of habitat they also have to respond to a dynamic landscape in which suitable
trees appear, grow and fall. Results show that, despite pre-supposed dispersal limitation
based on species’ fragmented distributions, many species exhibit high rates of gene flow
and little evidence of isolation by distance, suggesting that it may in fact be habitat
availability rather than dispersal limitation which constrains the distribution of these species
(Werth et al. 2006, 2007, Lättman et al. 2009). This is backed up by studies of bryophyte
colonisation on newly created habitats (e.g. slag heaps, Hutsemekers et al. 2008) and in
boreal forests (Hylander 2009) which show that many species are capable of long distance
dispersal, with a widespread, general ‘spore rain’ being the primary source of propagules,
enabling rapid movement across the landscape. Modeling of the effects of long distance
dispersal also suggests that species with this capability may be able to respond to rapid
changes in the distribution of habitats and would be relatively unhindered by the effects of
fragmentation (Pearson & Dawson 2005). Such rapid responses to habitat change, driven by
climate change, have already been documented in the Swiss bryoflora, where cryophilous
species have shown an upward range shift of 222m over the last 100 years, almost exactly
tracking the rate of temperature change (Bergamini et al. 2009). However, we might expect
that mountain species would be best able to track climate change, since the altitudinal shift
required to track a given temperature change would be a much smaller physical distance
than the equivalent latitudinal change.
2.3 Selecting target species
While some or maybe many cryptogam species are clearly capable of rapid responses in the
face of climate change impacts and other drivers, the difficulty for conservation will be to
identify those rare species which are limited by dispersal rather than by establishment or
habitat availability. Current conservation classifications use indicators such as range size,
occupancy and current population trends to assess threat status (e.g. Woods & Coppins
2003). Accurate prediction of which species are most threatened by climate change requires
knowledge of their autecology and dispersal ability; species with narrow niches but good
dispersal ability may not benefit from conservation action, we need to identify those species
with narrow habitat requirements and poor dispersal ability, where conservation action might
be most beneficial. In addition we need to identify species for which suitable habitats will still
be available under altered climate scenarios in the UK. In general, autecological information
for both common and rare cryptogam species is still scarce and this makes predictions
difficult.
2.4 AC as a tool for conservation
2.4.1 Transplantation methods
Experiments involving the transplantation of bryophytes and lichens within and between
localities have in the past been carried out for a number of purposes, particularly for pollution
monitoring, but also for studies of species’ ecological requirements and more recently as a
potential tool for species conservation. A number of different methods of transplantation
have been trialled, involving the movement of fragments of varying sizes. These range from
entire cushions or groups of shoots in bryophytes and multiple podetia or whole thalli in
lichens (e.g. Mitchell et al. 2004, Sonesson et al. 2007), through individual shoots/podetia, to
fragments of shoots and thalli (e.g. Hallingbäck 1990, Lidén et al. 2004, Gunnarsson &
Söderström 2007, Roturier et al. 2007). Many bryophyte and lichen species possess
specialised vegetative reproductive structures (see above); the use of these as a less
8

damaging (to the source population) source of propagules for transplantation has been
tested in lichens (Hallingbäck 1990, Scheidegger 1995), but not in bryophytes. Both
bryophytes and lichens also reproduce sexually by spores, but the ecological requirements
for germination of spores can be very precise (e.g. Sundberg & Rydin 2002) and spores
have not been tested as a means of AC.
One of the major issues with transplantation methods tested to date has been how to obtain
a successful physical bond between transplanted material and the recipient substrate. Most
bryophytes and lichens attach to their substrate by specialised structures such as rhizoids,
rhizines or anchoring hyphae and these are typically slow growing and may take some time
to sufficiently anchor a large adult individual. This issue is most problematic in epiphytic,
saxicolous and terricolous species of exposed habitats; in bryophytes of wet habitats e.g.
Sphagna, where shoots are packed together in a continuous carpet, it may not be a problem
(e.g. Gunnarsson & Söderström 2007). A number of transplant methods have been
developed to overcome the problem of getting good attachment of material to the new
substrate, these include transplanting propagules while still attached to their original
substrate. For example, bark flakes with epiphytic lichens have been glued onto new host
trees using epoxy resin (Hallingbäck 1990, Gilbert 2002a), complete twigs and branches
have been moved and fixed to a new host tree (Sigal & Nash 1983, Mitchell et al. 2004), and
pieces of rock have been moved with saxicolous lichens in situ (E. Coppins pers. comm.).
Where donor material has been removed from its original substrate it has been fixed to
recipient sites by a variety of methods including direct gluing of lichens with epoxy resin
(Richardson 1967, Gilbert 1988), tying on with string or thread (Lidén et al. 2004) or fixing in
place using a variety of types of mesh (Scheidegger 1995, Mitchell et al. 2004, Jansson et
al. 2009).
The majority of published studies concerning the use of transplantation to establish new
lichen populations or to reinforce existing populations have been carried out with epiphytic
lichen species, especially members of the lobarion community (Hallingbäck 1990,
Scheidegger 1995, Gilbert 2002a). Epiphytic transplants have been attempted by gluing on
of large thallus fragments, by rubbing of thalli onto the bark of potential host trees to lodge
small propagules in crevices, and by transplantation of soredia. Judging the long term
success of these transplants is difficult since few studies follow progress beyond 1-2 years,
but short-term survival rates of around 90-95% have been seen in Swedish forest lichens
(Lidén et al. 2004) when transplanting large fragments within the same site. Transplants to
new sites require a detailed knowledge of the species’ requirements and of conditions at the
potential host site, and were slightly less successful (85-90% survival). In one of the few
long-term studies, Gilbert (2002a) followed the progress of transplanted Lobaria amplissima
fragments over a 20 year period. Survival after 10 years was 70% but after 20 years this had
dropped to 40%; mollusc grazing, competition with bryophytes and detaching of the
transplanted substrate are all noted as causes of failure. Additionally, none of the transplants
gave rise to new colonies on the host trees, although several expanded beyond their initial
size. Transplantation of vegetative propagules and small fragments had much lower reported
establishment rates (Hallingbäck 1990, Scheidegger 1995); in the case of soredia the
survival rate after 2 years was around 10%, with most propagules which had established
after 2 months surviving in the longer term. However, these techniques allowed numerous
small thalli to become established with minimal damage to the donor population – an
important consideration for rare or endangered species (Scheidegger 1995). Thalli
established in this way are also likely to be less vulnerable to detachment of transplanted
substrates in the long term.
Transplantation of terricolous (growing on soil) lichens to assist colonisation of new areas
has also been reported in the literature. Roturier et al. (2007) tested the use of different sized
fragments of Cladonia mitis to re-establish lichen cover in boreal forest-floor habitats
following disturbance. They found that larger thallus fragments were more successful at
9

establishing, and also that the nature of the substrate played an important role in areas
where the transplanted lichens were exposed to wind; a mossy substrate proved more
suitable for lichen fragments to fasten to than twigs or bare soil.
There are relatively few reported cases in the literature of translocation being used as a tool
for species conservation in bryophytes. Transplantation has been tested for Sphagnum
angermanicum in Sweden (Gunnarsson & Söderström 2007) where whole plants and
fragments were moved within existing sites, to previously occupied sites and to new sites not
previously occupied but predicted to be suitable. Results showed success rates to be
highest with whole plants or large fragments and while there were large differences in
establishment rates between sites there was no overall difference between previously
occupied and new sites. For this species, where ecological requirements of pH and water
table depth are well characterised and typical associate species well known, suitable new
sites could clearly be identified with reasonable success. The habit of the species, growing
within an existing Sphagnum carpet, also meant that there were no issues with substrate
attachment. Bryophyte transplantation has also been used for re-introduction following local
extinction in the case of Scorpidium scorpioides in the Netherlands (Koojiman et al.1994).
Here, groups of S. scorpioides shoots were collected from Ireland and introduced to a Dutch
site where the species had previously been present. The transplant material was removed at
the end of the three year study, but by this time a good population of new shoots had begun
to establish up to 2m downstream of the transplants, suggesting that this could be a
successful method for this species if it was decided to go ahead with a permanent reintroduction.
2.4.2 Transplantation trials in the UK to date
Transplantation as a means of supplementing local populations or re-introduction to
previously occupied sites has been tested for a range of lichen species in the UK, including
several of conservation concern (A Coppins pers. comm.). The majority of trials have been
with epiphytic species, particularly members of the lobarion community as reported from the
literature (above). Epiphyte transplants have been tried using a variety of methods including
transplantation of bark sections with lichens in situ, gluing lichen thalli directly to a new host,
the ‘rubbing method’ of Hallingbäck (1990), stapling of lobes to a recipient tree under
sections of gauze similar to Scheidegger (1995), wedging of donor branches into the crown
of a recipient tree and ‘tucking’ of lichen fragments into bark cracks (used by Gilbert 2002b)
for Teloschistes flavicans). Results have been very variable, with several transplants failing
in the longer term due to mollusc grazing, though apparent successes have been achieved
with Lobaria pulmonaria, L. amplissima, Sticta limbata and Teloschistes flavicans. Best
successes were noted for transplants into the tree canopy rather than onto the main trunk,
possibly due to reductions in slug predation. Translocation has also been attempted for
terricolous species on the Breckland heaths (Fulgensia fulgens, Buellia asterella and
Squamarina lentigera) but all three species failed to thrive at their new locations.
Transplantation has also been attempted in the UK for a small number of threatened
bryophyte species including Bryum muehlenbeckii, Bryum schleicheri and Orthodontium
gracile. With B. muehlenbeckii (a cushion forming species of riparian habitats) material was
transplanted as small ‘plugs’ of shoots held in place at the recipient location by cocktail
sticks, or alternatively by sewing shoots into small muslin bags which were then glued into
place, allowing the moss to send axillary shoots up through the material. This latter
technique was also tested for B. schleicheri, a spring species (Rothero et al. 2006) with
material derived from in vitro cultivation. Results have been mixed with between 20 and 40%
of transplants surviving a 1-2 year period after transplantation (G. Rothero pers. comm.)
however, both are ‘weedy’ species occupying fairly dynamic habitats and there is some
evidence to suggest that the transplants have led to establishment of populations outside of
the initial transplant material.
10

2.4.3 Risks associated with AC
It can be seen from the above discussion that, aside from the methodological challenges of
achieving successful transplants, many fail due to our limited ability to select suitable host
sites. It is important that we address this issue before attempting to use AC as a large scale
conservation technique for rare species, as otherwise we run the risk of unnecessarily
depleting source populations.
The risk of introduced species becoming invasive is also a frequently cited problem with the
process of AC. Many bryophyte and lichen species have a high dispersal capacity, as
demonstrated by their role as primary colonisers in dynamically changing habitats
(Söderström 1992) and by the rapid range expansions and contraction they may display in
response to anthropogenic drivers such as air pollution (Söderström 1992, Gilbert 1992,
Massara et al. 2009). Despite, or perhaps as a result of this capacity for natural long
distance dispersal, there are few reported instances of lichens or bryophytes becoming
invasive. Lichens and bryophytes have rarely been deliberately introduced, but have
occasionally been accidentally imported to new areas, frequently with vascular plant
specimens for gardens. Twenty-two species of bryophyte had been identified as naturalised
aliens in Europe by 1992 (Söderström 1992) and of these only 3 were invasive
(Orthodontium lineare, Campylopus introflexus and Riccia rhenana) with a fourth,
Lophocolea semiteres, arriving more recently (Hassel & Söderström 2005). The majority of
these are southern hemisphere species and, despite undergoing a rapid spread, have not
caused any major environmental problems, with the possible exception of C. introflexus,
which may cause problems in heathland and dune ecosystems. Those species which have
become invasive are generally characterised by abundant spore production and/or
production of vegetative propagules which enable rapid spread and colonisation of available
habitat; it seems unlikely therefore, that species being considered for AC to enable short to
medium range dispersal would share the attributes of these invasive species, since they
would otherwise be able to exploit a dynamically shifting habitat without assistance.
2.4.4 What is required for successful AC?
From the above discussion, three main requirements for achieving successful AC can be
determined. Firstly a means to identify which species are most at risk from climate change,
i.e. those which are dispersal-limited. For non dispersal-limited species, maintenance of
healthy functioning ecosystems that provide appropriate niches should be the focus.
Secondly, assuming we can identify those species we wish to move, accurate description of
niche requirements is needed, for both adult and juvenile stages, including knowledge of the
interactions with predators such as molluscs. This would allow more accurate selection of
suitable host sites (both now and for the future) and increase the likelihood of establishing a
viable population. Thirdly, effective methods for the transplantation of adult individuals or,
preferably, sexual or asexual propagules are needed to achieve a transplant that is
successful in the long term.
2.5 Summary & suggestions for experimental studies
Currently there is a mismatch between the habitats in which we have some knowledge of
species’ dispersal abilities and landscape scale dynamics (habitats with a high proportion of
epiphytes) and those which appear likely to be most impacted by climate change (arcticalpine
habitats). Investigations of dispersal abilities and genetics in arctic-alpine species
would be very useful to determine if certain groups of species would benefit from assisted
colonisation.
Assuming that species in need of help could be identified, practical trials of transplant
methodologies for terricolous and saxicolous species would be of value, since most work to
date has focused on epiphytes. These should probably focus on use of spores and
11

vegetative fragments rather than adult thalli if possible and should be carried out with
properly replicated, statistically sound methodologies and long term follow up, which have
generally been lacking to date.
12

3 SPECIES GROUP NO. 2 – VASCULAR PLANTS
3.1 Threats to vascular plants in Scotland
In Scotland, we might initially assume that AC would benefit species that:
1. Will have no climatically suitable environments within their current range.
2. Are unable to track suitable climate unaided.
3. Have suitable climate space in Scotland under future climate change scenarios.
4. Have poor capacity to adapt in situ.
Although these may appear relatively simple criteria, it may be much more complex to
determine which species might be suitable candidates.
3.1.1 Climate and the ranges of Scottish vascular plants
The Scottish vascular flora includes (for the UK) a high proportion of species from particular
“geographical elements” as designated by the phytogeographer J.R. Matthews (Lusby &
Wright 2001). These species commonly have their southern or low-altitude range limits
within Scotland, and include members of the arctic-subarctic element (species absent from
central Europe), the arctic-alpine element (often circumpolar with main distribution focused
on the Arctic, but occurring in some more southern, alpine locations), and a smaller group of
species within the Alpine element (occur in central European mountains, but absent from
Northern Europe). Very few of these species have a significant proportion of their population
in Scotland and so, at a global scale, the future survival of the Scottish populations is not of
absolute conservation concern. However, within the Scottish context, and recognising the
obligation under the CBD for conservation of species at the national level, these species
probably represent some of those most obviously threatened by future climate change. UK
Biodiversity Action Plan (BAP) species for which climate change was identified as a
substantial threat include, for example, mountain scurvy grass Cochlearia micacea,
Newman’s Lady-fern Athyrium flexile (although their taxonomic status is debatable), and
Norwegian mugwort Artemisia norvegica (Brooker et al. 2004). The last of these is noted as
an exception to the “common elsewhere” rule, as it only occurs in central and southern
Norway, and the Ural mountains in Russia (Lusby & Wright 2001).
The direct impact of climate on plants in unproductive and harsh arctic-alpine environments
is proportionally much stronger than in more productive systems at lower altitude or latitude
(Brooker 2006). The lower latitudinal or altitudinal limits of these arctic-alpine species are
therefore likely to be strongly influenced by climate, and to respond (relatively) directly to
climate change through northward or upward contraction. Such theoretical predictions of the
strong climate-distribution link in northerly species are supported by the recent assessment
by Beale et al. (2008) of climate envelope modeling approaches (for details, see Section 5).
However, given the highly restricted and often high-altitude distributions of such species, it is
questionable whether suitable climate space will be available for them in Scotland following
future climate change. In addition, even for arctic-alpine species, where the climatedistribution
link is considered to be much tighter, southern and low-altitude range limits may
be under less direct climatic control than northern or high-altitude limits. Range limits in more
productive systems tend to be more strongly regulated by interactions with other species,
rather than directly by climate (Vetaas 2002, Brooker 2006)
Despite the potential complexity, though, it does appear that mountain-top vascular plant
species in Scotland may be responding to climate change in the expected manner. A recent
study by Britton et al. (2009) involved revisiting a widespread set of botanical plots initially
recorded between 1963 and 1987. They found a general increase in species richness over
time across all sites, but that this was stronger for vascular plants and bryophytes than for
lichens. Overall there was a general homogenisation in the composition of plant
13

communities, possibly due to expansion of widespread generalist species in response to
warming as well as to nitrogen deposition. The sizable changes in snowbed systems, as well
as the decline in cover and frequency of species with northern and alpine distributions,
indicate a climate impact.
For other species with the southern limits of their distributions in Scotland, it is less certain
that these limits will respond to climate change. In more productive Scottish systems, at
lower altitude or latitude, land use and habitat fragmentation become stronger regulators of
species’ distributions and abundance. Vascular plant species comprising part of the Boreal
element of European vegetation (Polunin & Walters 1985) have southern range margins
within Scotland, but are commonly found at lower altitudes than arctic-alpine species. They
are thus more likely to have suitable climatic conditions in Scotland under future climate
change, but their distributions have often been more strongly influenced by anthropogenic
habitat fragmentation than arctic-alpine species. Thus the strength of the link between their
current realised Scottish “warm” range limit, and climate, may be weaker. For example,
twinflower Linnaea borealis, although being a species judged to be “probably affected” by
climate change (Brooker et al. 2004), has a Scottish distribution which is strongly influenced
by its affinity to Scots pine woodland. It is therefore difficult to judge whether this southern
range limit currently reflects an absolute climate tolerance (either direct or bioticallymediated),
or sits within a climate envelope that extends further south than the current
realised range. One-flowered wintergreen Moneses uniflora (another pinewood specialist) or
the maritime oysterplant Mertensia maritima (which is strongly affected by coastal grazing
(Lusby & Wright 2001)) may represent similar challenges in predicting the impact of climate
change, although for the latter absolute physiological tolerances have been suggested to set
southern range limits (Crawford 1989). Such uncertainties highlight the lack of long-term
population-level monitoring data (i.e. abundance and success of populations rather than
nation-wide distributional data) for many rare plant species. Such data has been critical in
studies of bird species with respect to detecting and understanding the impacts of climate
change (e.g. Moss et al. 2001, Beale et al. 2006). Furthermore, the inevitably limited
distribution data for rare species makes the fitting of climate envelope models more
problematic (see Section 5). Hence, for rare species it might be much harder to prove
definitively that climate change is driving distributional or abundance declines, and that
extinction of populations or total loss of current range is imminent and thus that AC is the
only remaining option.
3.1.2 Plant dispersal and habitat fragmentation
One group of vascular plants that might clearly benefit from AC are those with limited
capacity for dispersal. Several factors can limit vascular plant dispersal ability. Some species
have evolved characteristics that enable them to disperse over longer distances, e.g. small
seeds, released at a greater height on the plant (increasing the chances of encountering
some kind of airflow), and with adaptations to improve “buoyancy” such as the feathery
pappus on the seed of thistles and dandelions. Other species, in contrast produce far fewer,
much heavier seed, which fall mainly in the immediate vicinity of the parent plant. In a static
environment this latter group benefit from a greater chance of any given seed landing in
suitable habitat, which is why a seed size vs. seed number trade-off is thought to occur
(Grime 1979). However, producing a few heavy seeds is thought to severely limit a plant
species’ long-distance dispersal abilities. At the same time many of the plants that produce
such seed are also relatively long-lived, often perennial species. In contrast multiple small
seeds are characteristic of classic disturbance-tolerant “ruderal” species such as annual and
biennial arable weeds. Large-seeded plants therefore not only produce fewer seeds per
plant, but it may take longer for the plant to reach reproductive maturity, further limiting the
number of seeds produced per year. We would thus expect that large perennial plants,
producing few large seeds per year, may have much slower dispersal rates than small
annual plants producing many (often wind-dispersed) seeds. The exception to this would be
where seeds are zoochorous, i.e. animal dispersed, in which case large seeds can be
14

transported over large distances, although this would not increase the absolute rate of seed
production which may continue to limit dispersal rates (as opposed to distances).
Unfortunately, empirical data on the realised dispersal distances of vascular plants is
relatively limited, not least because of the difficulty of measuring the highly infrequent longdistance
dispersal events that appear to be critical in determining plant dispersal (Cain et al.
2000). Those studies that do exist would indicate that expected trait-dispersal relationships
may not always hold true. For example Honnay et al. (2002) found that the impacts of
fragmentation on colonisation of woodland fragments by woodland-specialist plant species
depended on seed dispersal mode. The best dispersers were endozoochores (dispersed by
ingestion by birds and mammals) and epizoochores (dispersed by adherence to birds and
mammals). However, Honnay et al.’s results do not entirely match naïve expectations - the
worst colonisers were anemochoric (i.e. wind-dispersed). For such species, although their
propagules may disperse further, the probability of their arrival in suitable habitat for
germination and growth is actually much lower than for animal-dispersed seeds. Genetic
techniques - in particular the application of microsatellite markers - are now also enabling us
to assess directly the level of gene flow between individuals and populations, and to explore
whether expected relationships between gross morphological traits and gene flow hold true.
As shown in Honnay et al.’s work, it appears that naïve expectations are not always
confirmed. For example, Wilk et al. (2009) showed a high degree of genetic isolation
between spatially-close populations of Fragaria virginiana, despite insect pollination and
endozoochorous seed dispersal. Similarly Scobie & Wilcock (2009) found considerable
genetic isolation of twinflower Linnaea borealis populations, even over very short distances.
Such results also highlight the second main factor that can regulate plant dispersal: the level
of habitat fragmentation. Scobie & Wilcock (2009) found a significant reduction in
colonisation success in a highly fragmented landscape compared to a more connected one.
This is again related to the availability of suitable environmental conditions, either for the
germination and growth of the plant, or – in the case of zoochorous species - for its animal
vector. Dispersal-limited species may therefore be those with either poor inherent dispersal
abilities, or with requirements for habitats which are themselves highly fragmented. This may
in part explain why, in those cases where changes in plant species distributions have been
observed, it tends to be widespread generalist species whose ranges are expanding
(Preston et al. 2002). Importantly, for some plant species, climate change may directly act to
increase habitat fragmentation. For alpine species, range shifts to increasing altitude are
also associated with increasing distance to other mountain-top areas which may contain -
now or in the future - suitable habitat (Hester & Brooker 2007). Coastal systems may also be
affected – changes in storm patterns and associated changes in shoreline processes may
alter the distribution of suitable coastal habitat for species such as Mertensia.
3.1.3 In situ adaptation to climate change
Even for those vascular plant species where there is likely to be an impact of climate on the
current range, and where dispersal is likely to be problematic, climate change might not
necessarily constitute a threat if such species are able to adapt to climate change in situ.
Such evolutionary shifts in physiological tolerance might involve, for example, changes in the
temperature relationships of some physiological processes. Some arctic species such as
Mertensia or Scots lovage Ligusticum scoticum have a relatively high rate of respiration,
which in turn responds rapidly to changes in temperature (Crawford 1989). Hypothetically,
evolutionary changes might reduce its temperature responsiveness, or the temperature at
which it reaches a maximum. In terms of indirect climate effects, competition is clearly a
major regulatory factor at southern range limits (Brooker 2006). Evolutionary responses
might therefore involve greater competitiveness or shifts in phenology to avoid periods of
strong competition. Evolution may even enable dispersal limitation to be overcome, as
shown for both invasive plant species (Davis 2005) and at the expanding range margins of
insect species responding to climate change (Thomas et al. 2001).
15

However, evidence of in situ adaptation, when compared to evidence of range shifting, is
extremely limited. A palaeoecological study by Davis et al. (2005) has shown that inclusion
of evolutionary processes is likely to be important in explaining anomalies in historical (i.e.
post glaciation) range shifting data and hence that in situ adaptation has probably occurred
during previous climate change. However, it is important to note substantial differences
between post-glacial and current climate change in terms of rate. The fast rate of current
climate change may limit the evolutionary potential of many species, particularly those with
long life-spans and times to reproduction. Other factors limiting evolutionary potential are
depauperate within-species genetic diversity (limiting the overall gene pool from which “fit”
genes can be selected), and limitations on gene flow across species ranges (Willis et al.
2009). These latter factors may be particularly important for rare species, with fragmented
and small populations which may have undergone genetic depauperation and which,
consequently, can suffer substantial difficulties in terms of exchanging and recombining
genes between individuals (Barrett & Kohn, 1991, Ellstrand & Elam 1993, Schaal & Leverich
1996, Young et al. 1996, Wilcock & Neiland 2002). Furthermore, the reproductive problems
stemming from small population size can in turn reduce the output of viable seed (Willi et al.
2005, Honnay et al. 2006, Scobie & Wilcock 2009), which further limits dispersal potential.
3.2 Possible risks from assisted colonisation
Vascular plants can clearly become invasive, as demonstrated by some of the examples
given in Section 1. Notably, cross species-group comparisons have found that there is
relatively less risk from vascular plant invasions compared to other species groups such as
freshwater vertebrates (Mueller & Hellmann 2008) although, and again as noted (Section 1),
this does not mean that there is no risk. Increasing evolutionary distance between the
introduced species and recipient community increases the risk of an introduced plant
becoming invasive (Strauss et al. 2006). Invasive plant species can have detrimental
consequences for recipient communities by direct or indirect competitive exclusion of
existing species (Levine et al. 2003), disrupting disturbance regimes such as fire cycles
(D’Antonio 2000), or disrupting plant-insect (e.g. pollinator) mutualisms within the native
community (Traveset & Richardson 2006).
In the Scottish context the risk of invasion of recipient communities by AC target plant
species might be considered relatively low. Some species with the potential for AC (e.g.
arctic-alpine species) are inherently slow-growing and uncompetitive. Furthermore we might
argue that rare species are inherently unlikely to invade, and as AC would be within
Scotland, the likelihood of getting substantial evolutionary differences between the source
and recipient communities might be considered low. However, all of these assessments are
relative: we cannot be certain, because of the currently unpredictable nature of such events,
that there is zero risk of an AC vascular plant species becoming invasive.
Similar uncertainty surrounds the possible impact of translocated vascular plants on
ecosystem function. Impacts might include, for example, increased productivity within a
system following the deposition of more readily-decomposed litter, as documented following
the expansion of birch woodland onto moorland systems (Mitchell et al. 2007). Although AC
species may be subdominant within communities, the scale of impact is not necessarily
related to biomass. Subdominant hemi-parasitic plants, for example, have been shown to
have substantial impacts on nutrient turnover processes within subarctic dwarf shrub heath
communities (Quested et al. 2003, Watson 2009). Furthermore, AC species will not
necessarily remain subdominant in the recipient community, and our lack of understanding
of the regulation of ecosystem processes by plant traits (Lavorel & Garnier 2002) makes the
possible impacts of new plants, with potentially unique traits, hard to predict.
16

A risk from AC which may be greater for vascular plants than for other species groups is
hybridisation. In the UK this differs greatly between plant taxa. Some genera (e.g. members
of the Salicaceae or Cyperaceae) clearly hybridise readily. However, based on extensive
knowledge of existing UK hybrids, such as that provided by Kent’s lists of UK vascular plants
(Kent 1991), or in the pending output from the current BSBI hybrids project, it is possible to
assess whether a species readily hybridises, and also whether species within the recipient
community might be cross-compatible with it. Notably, hybridisation is only very infrequently
a mechanism for the detrimental impact of invasives (Gurevitch & Padilla 2004a).
Hybridisation might therefore represent a small element of the risk from AC.
3.3 Factors determining the practicality and success of assisted colonisation
3.3.1 Propagule type
There have been a number of studies trying to assess the factors that determine the success
of translocations of vascular plants, as well as the practicality of different translocation
options (e.g. Maunder 1992, Menges 2008, Soorae 2008, Dalrymple et al. In Prep). Many
translocations of vascular plants have been undertaken, either as re-introductions, or in
some cases as genuine AC when the original native habitat was destroyed or restricted to a
single remaining site (for examples see Maunder 1992, Swarts & Dixon 2009). However, a
consistent comment throughout these reviews is that the lack both of careful recording (of
procedures and data) and of long-term monitoring makes it difficult to assess whether and
why a translocation has been successful. However, in those cases where the latter is
possible, a number of issues have been identified as being particularly important.
Propagule selection plays an important role. Several studies note that adding seed of the
target species to the recipient site is much less effective than transplanting juvenile or adult
plants (Maunder 1992, Menges 2008, Dalrymple et al. In Prep). The latter, obviously, are
more expensive in terms of collection, transport and propagation, but have much lower
mortality and overcome the problems associated with seed dormancy mechanisms, or at
least allow such mechanisms to be specifically dealt with under controlled conditions (e.g.
Krauss et al. 2002, Maschinski & Duquesnel 2007). Furthermore the pattern of introduction
into the recipient site can be more effectively controlled, as there is less chance that a low
proportion of establishment (as found with seed) will lead to spatial patterning in the new
population that limits success. For example, Colas et al. (2008) created new populations of a
cliff-dwelling Mediterranean endemic Centaurea corymbosa. In the new populations there
was reduced fecundity because seed-sowing produced a low density of suitable crosses in
this self-incompatible species. However, they also found that the new populations had higher
survival rates, leading to similar asymptotic growth rates in the donor and new populations.
In the Scottish context it may be difficult to get successful translocation via seed, particularly
for arctic-alpine species which tend to propagate clonally and produce only a few seed which
are often infertile (Callaghan et al. 1997). In contrast propagation of clonal tillers can be
relatively straightforward, particularly for graminoid species (R. Brooker, pers. obs.).
3.3.2 Repeated propagule additions and selection of suitable sites
High mortalities can sometimes necessitate repeated introductions over several years
(Maunder 1992). This has been proposed as a mechanism to overcome, for example, limited
mate availability (Colas et al. 2008). However, repeated introductions can also benefit from
ongoing monitoring and an adaptive management approach. For example, Maschinski and
Duquesnel (2007) found that some of the introduction sites for Sargent's cherry palm
Pseudophoenix sargentii in the Florida Keys had very high levels of mortality, possibly as a
result of exposure to storm surges. Subsequent introductions were able to avoid this
particular microhabitat. Dalrymple et al. (2008) found that initial differences in germination
success between sowing sites for small cow-wheat Melampyrum sylvaticum occurred
because of differences in seed dormancy – once this had been accounted for there were no
17

discernible differences in recipient site suitability or donor population fitness. In this case
long-term monitoring may have prevented new additions of seed when they were, in fact,
unnecessary. Similarly, highly detailed long-term monitoring, in combination with population
viability analysis, can help identify factors that are limiting the success of plants at an
introduction site, and corrective management actions can then be put in place (Colas et al.
2008). Such adaptive management can also help overcome the difficulty of precisely locating
the environmental conditions necessary for population establishment. It has been suggested
that the recipient site should match, as closely as possible, the conditions found at the
current location (Swarts & Dixon 2009). However, as pointed out by Maunder (1992) “If a
species from a degraded or relict habitat has not reproduced for decades how can the
regeneration niche… be identified? The field or collection notes on the distribution of mature
plants are not necessarily informative about the site for reintroduction …”. Furthermore, if we
are planning to move target species into areas that will be suitable under future climate,
should we expect the vegetation within these areas to currently match that of the donor
population?
3.3.3 Genetic constitution of propagules and problems of ex situ propagation
As well as the type of propagule to be used, and the question of locating a suitable site, the
genetic diversity of introduced stock can also be important. As shown by the work of Colas et
al. (2008) limited genetic diversity can lead to problems of reduced fecundity, either through
self-incompatibility mechanisms or inbreeding effects following fertilisation, and is already a
problem for many species of rare vascular plant (e.g. Oostermeijer et al. 1995, Luijten et al.
1996, Fischer et al. 2003, Scobie & Wilcock 2009). Unfortunately, given the small population
size of those species that might be considered for AC, the donor population may already be
genetically depauperate (Maunder 1992). In addition, larger populations are better able to
cope with seed removal (Dalrymple et al. 2008). This is why some authors recommend
undertaking AC prior to it becoming a tool of “last resort”, or at least recognising that the
point of “last resort” relates to much larger populations than we might naïvely suppose.
Genetic depauperation might be overcome by introducing propagules from multiple
populations into a single recipient site, but in turn this raises the alternative risk of
outbreeding depression and AC failure (Menges 2008), although this will clearly not harm the
donor populations themselves. An additional problem is the selection of donor populations –
is it better, when considering populations to cope with the future impacts of climate change,
to select individuals from the “warm” range margin, which would therefore, possibly, be preadapted
to warmer temperatures, or is it better to select from across a species range in
order to provide maximum adaptive potential for a range of factors, including climate? It is
extremely difficult to answer these questions, not least because firstly we have, for many
species, very limited knowledge of the distribution and function of genetic diversity in wild
populations. For example, we might expect some species to be genetically depauperate
because of their low population size, but other processes, for example low rates of
reproduction and long life-spans can maintain genetic diversity even in small populations.
Generally, populations of long-lived (and especially clonal plants) suffer a gradual loss of
genetic diversity under reduced or completely absent sexual reproduction as the fittest
genotypes or clones out-compete the others (see Schaal & Leverich 1996). The length of
time that a population has been small (and remains small) is a critical factor concerning
losses of genetic diversity. This can be heavily influenced by management history
(Jacquemyn et al. 2009), but again our knowledge of local population dynamics – rather than
distributional changes - is likely to be limited for many rare plants. In addition, would it be
feasible, for species at the point of “last resort”, to undertake a thorough population genetics
analysis and study of genetic issues associated with long-term reproductive success? Along
with the need for long-term monitoring, as well as potentially expensive ex situ propagation,
substantial sums of money would be required for a thorough genetic survey of a species.
Ex situ propagation of target species may have negative impacts on the genetic diversity of
introduced populations (Maunder 1992). Other problems may also result from ex situ
18

propagation, including the acquisition of diseases which are then transported, along with
plant propagules, to the recipient site (Hodder & Bullock 1997), or the production of inbred
seeds which survive in the nursery but not in the wild (Kettle et al. 2008) However
procedures can be put in place to address problems such as disease transfer (e.g. Kraus et
al. 2002), or the collection of genetically impoverished material prior to ex situ propagation
(Kettle et al. 2008). Indeed the IUCN guidelines specify the need to quarantine translocated
organisms, particularly those from ex situ collections (IUCN 1995), although this may not be
straightforward. Interestingly, though, stripping a species of its pests may have negative
consequences for the recipient community because some pests or pathogens may help to
prevent a species from becoming invasive. For example, the build up of native soil
pathogens is important in driving successional replacement of the dominant Ammophila
arenaria in dune systems, and their absence may be the cause of Ammophila’s invasive
dominance in areas to which it has been introduced (Van der Putten 2009).
3.3.4 Recreating essential interactions
The re-creation of essential interactions in the recipient location is not just an issue with
respect to the regulation of invasiveness. Many plant species are dependent on mutualisms
with a wide range of other taxa, including insects and fungi (Traveset & Richardson 2006). In
some cases the role of a missing mutualist might be performed by alternative species in the
recipient location, for example by other generalist insect pollinators. However, some plant
species appear to have very tightly co-evolved relationships with pollinators (Glover 2009).
Seed dispersal activity may also be limited by the absence of specialist mutualists, for
example seed-foraging ants. Essential interactions also involve intra-specific effects.
Introductions may therefore also need to consider meta-population dynamics relevant to the
AC species – do multiple sub-populations need to be created, along with vacant but suitable
space, for essential meta-population processes to occur?
3.3.5 Application of guidelines
Many of these issues are covered by relevant guidelines, for example those of the IUCN
(1995). However, it is questionable how often and how strictly these guidelines are followed.
For example, even within the IUCN’s own recent publication on re-introductions (Soorae
2008), it seems that very few of the case studies presented actually follow these guidelines
in full. This is probably not because of a lack of commitment on the part of those individuals
undertaking re-introductions, but because of the complexity of following all of the steps
proposed in the guidelines, as discussed in Section 7.2. In addition, adherence to these
guidelines is not a guarantee for translocation success (Dalrymple et al. In Prep.).
3.3.6 Other practicalities: monitoring success, historic ranges, and alternatives to AC
Beyond factors regulating practicality and success, several other important points have been
made in recent reviews of plant translocations. Firstly long-term monitoring, although not
necessarily and directly impinging on the success of a given translocation, is something that
is strongly encouraged. Ultimate success of a translocation can be defined as the
establishment of a self-sustaining population which is larger than that originally introduced to
a site. Assessing this, especially for slow-growing, long-lived, and clonal plant species might
be particularly difficult, and population size might itself be assessed using a number of
metrics (e.g. more free-living individuals, more genetic diversity, or more heterozygosity).
However, a number of interim measures of success have been proposed, including seed
germination, plant growth rates, time to reproduction, flowering levels and seed production
and, as noted above, these various parameters can also be used for population viability
analysis to assess the translocations’ likely long-term success (Menges, 2008). An important
factor to include in long-term monitoring is the monitoring of both donor and recipient
populations and communities, as well as monitoring of undisturbed source populations and
empty potential recipient sites. This would allow the impact of the act of translocation on the
source and recipient systems to be assessed and isolated from the impact of concomitant
but unrelated factors. In other words, it would enable us to see whether observed changes in
19

the donor or recipient systems would have occurred in the absence of AC. This will be
essential for future, objective, evidence-based assessments of AC.
A second key issue is that of determining whether or not a site is within a species’ historic
range. This is critical in terms of determining whether or not a translocation is an AC or a reintroduction,
and hence which set of guidelines and principles apply. As discussed later
(Section 7) the JNCC guidelines take a threshold date of 1600 as the cut-off for current
ranges, but meta-population dynamics and the difficulty of precise recording for infrequent
and rare species in isolated areas mean that it is very difficult to be sure that a location is - or
is not - within a species’ historic range.
Finally, given the complexity of the issues, and the difficulty of following the published
guidelines, it has been suggested that ex situ collections should be the primary focus of
activity, as they enable species to be conserved until the underlying information necessary
for a successful and safe AC has been acquired (Armstrong & Seddon 2008, Vitt et al.
2009). However, and as pointed out above, ex situ propagation, especially when coupled
with uninformed collection of material for propagation, can have unexpected and detrimental
impacts on the target species, not least in terms of the genetic structure of the propagated
material. Furthermore, at what point do we start putting the conserved species back into the
wild? This question is going to become increasingly difficult to answer if natural systems do
start to respond rapidly to climate change. The existing native communities may no longer
exist, and we would then still be faced with the same problem of selecting a potential
recipient system.
3.4 Summary & suggestions for Experimental Studies
In summary, although we can list characteristics that we might consider provide us with a
useful “first sift” of vascular plant species likely to benefit from AC, there is considerable
complexity in terms of making further choices. Those species whose ranges are clearly
climate-restricted, and are likely to respond directly and negatively to climate change, may
also be those least likely to have suitable climate space in Scotland, i.e. arctic-alpine
species, and hence AC is simply not an option. For those species with greater potential for
the occurrence of suitable climatic conditions, we also see a greater current impact of habitat
fragmentation, and so reduced certainty that current distributions are climate-limited. At the
same time, if these species are climate-limited, they will clearly suffer - because of habitat
fragmentation - in terms of reduced capacity to track suitable climatic conditions or for
evolutionary adaptation in situ. Although many rare plant species might find it very difficult to
cope with climate change - making them good candidates for AC - these same species
present the greatest difficulties in terms of predicting future negative climate impacts, and
the availability and location of suitable climate space. Unfortunately it is for these most
threatened species that we might need to act most rapidly, either through ex situ
conservation or AC prior to further genetic loss, in order to ensure that AC is successful.
In terms of risk from AC, there are clearly risks from invasion, with numerous possible
impacts on recipient communities. These risks are likely to be low for vascular plants,
especially when translocated to sites near to their donor communities in both physical and
evolutionary space, and particularly in relatively unproductive environments. However, these
risks exist, even if at a low level.
In terms of the practicalities of undertaking AC a large number of factors need to be
considered, including impacts on the source population, genetic structure of translocated
material, form of translocated material, need for and possible impacts of ex situ propagation,
the occurrence of essential biotic interactions, and the selection of suitable recipient sites.
Clearly the common recommendations of a thorough knowledge of the genetic structure of
20

translocated material, the likely impacts of AC on source and recipient sites, and the need
for long-term monitoring and possible repeated introductions, all make AC a potentially
costly activity. It may be, though, that standardised approaches based on sound theoretical
principles could avoid the need for perfect understanding of the target species. For example,
in terms of dealing with potential genetic problems during AC, procedures such as the
selection of large donor populations, the careful sampling of multiple individuals as donors,
and the tracking of success of transplanted individuals during the early stages of AC to avoid
cryptic bottlenecks may consistently overcome such problems.
However, many of the recommendations regarding AC are based on a relatively limited
knowledge-base, not least because of inadequate recording and reporting of previous
efforts. Targets for research to help expand this knowledge base include:
Assessing which species are likely to be candidates for AC: possibly using a risk
assessment process that combines trait data (e.g. for dispersal ability, reproductive
characters, growth form) to assess likely impacts of climate change and habitat
fragmentation - along with risk of invasiveness - with modeling data on likely
occurrence of suitable future climate space and/or potential risk of decreasing
landscape connectivity from further habitat fragmentation.
Understanding how to assess climate impacts on species with heavily fragmented
distributions.
Better population-level monitoring of rare plant species so that climate-driven
declines can be assessed (or decline assigned to some other environmental driver),
helping with the risk assessment process.
Better understanding of what constitutes “last resort” for rare plant species - how is
this influenced by e.g. breeding systems?
Exploring the best locations from which to source material for AC relative to current
and future climatic conditions.
Combining climate projections and knowledge of vegetation transitions to select sites
that will be suitable for future survival of the species. Should we select recipient sites
that look like the current habitat of the species?
21

4 SPECIES GROUP NO. 3 – TERRESTRIAL INVERTEBRATES
4.1 Threats to terrestrial invertebrates in Scotland
At least 24000 invertebrate species are known from Scotland, of which over 1400 are not
found elsewhere within the UK (Macadam & Rotheray 2009). More montane and boreal
(north European) species exist in Scotland than elsewhere in the UK. Scotland is also the
last stronghold for an increasing number of invertebrate species that have become rare or
extinct elsewhere (Macadam & Rotheray 2009).
Five terrestrial invertebrate species (and many more marine invertebrates) are entirely
endemic to Scotland. Whilst few Scottish invertebrates receive any sort of legal protection,
there are 166 UK BAP invertebrate species that are found in Scotland, and 350 invertebrate
species are on the Scottish Biodiversity List (Macadam & Rotheray 2009).
As for vascular plants and lichens, defining which invertebrates are threatened but may be
aided by AC is difficult. Climate-driven changes to insect distributions may be hard to detect
as accurate data are lacking for many species groups. Long-term data for Lepidoptera exist
in some European countries including, for example, the UK Butterfly Monitoring Scheme
(see, e.g. Asher et al. 2001). There is also the Rothamsted Insect Survey network of light
traps in the UK (Woiwod 1997) although this latter dataset is thought to contain many errors.
Elsewhere, data on former and current insect distributions is far more sparse (New 2008).
4.1.1 Dispersal and habitat fragmentation
Some upland/northern UK butterflies have disappeared from lower altitudes (Hill et al. 2002).
This may be a response to either climate or land use change. However, where suitable
habitat still exists on the same mountain, distances are likely to be within the dispersal
capability of most butterflies. A greater threat exists for species occupying the top of a
mountain or an otherwise isolated habitat patch. An example in Scotland is the moth
Mountain Burnet (Zygaena exulans subochracea) which, despite feeding primarily on a
widespread plant, Crowberry (Empetrum nigrum), is restricted to small areas on a couple of
high mountains near Braemar. New (2008) considered criteria under which Australian
butterflies may be assessed as candidates for assisted colonisation. Z.e. subochracea might
be considered analogous to category 2 of New’s scheme: i.e. species with a single
population in a presumed intermediate part of the potential range, where there is also a
presumed threat from climate change (such as for mountain-top dwellers or species in
otherwise isolated habitat patches). In cases such as this moth, AC in a poleward direction
may be the only conservation action available. New also considers AC to be an important
tool for conservation of species found in a few isolated populations, but again where sites
suitable for natural poleward dispersal are unavailable. However, and as New points out,
recipient sites might need to be prepared years or decades in advance, and the resources
and political will for such operations might not exist for invertebrate species.
4.1.2 Habitat specialism and local adaptation
In a changing environment, habitat specialists are under greater threat than generalists. This
leads to a trend, similar to that noted for vascular plants, of replacement in assemblages of
specialist species by those able to utilise a broader range of resources (Polus et al. 2007).
Habitat specialists may have a more limited dispersal ability than generalists (e.g. Braschler
& Hill 2007, Littlewood et al. 2009) and behavioural traits, such as having a home range,
may further limit their ability to disperse even within a habitat patch (Korosi et al. 2008).
Local adaptation, in particular to climatic conditions, may also limit dispersal. Reciprocal
translocations between central and edge (poleward) populations of a host-plant specialist
butterfly show that local adaptations by peripheral populations to low winter temperatures
may inhibit poleward movement under warming conditions (Pelini et al. 2009). Reciprocal
22

transplants have also shown that phylogenetically similar species may respond differently to
varying conditions at their range boundaries. For some species the boundary edge may be
only marginally suitable in terms of climate, and hence warming conditions may increase the
suitability of such areas and extend the area of climatically suitable conditions. Other species
may show less response at the range edge, especially those species for which range
boundaries are dictated largely by factors other than climate. Hence apparently similar
species in the same area may show differing capacities for poleward expansion (Hellmann et
al. 2008).
Overall, the distributions of habitat generalist species are more closely related to climate
than are the distributions of habitat specialist species, which instead are more closely linked
to host-plant abundance (Menéndez et al. 2007). Whilst for some common generalist
species there is ample evidence of a general northerly spread in distribution (e.g. Parmesan
et al. 1999), dispersal limitation through lack of suitable habitat has been identified in UK
butterflies, with 30 out of 35 species studied by Hill et al. (2002) failing to keep track of
recent climate changes. Even some species traditionally regarded as generalist may be
struggling to utilise all climate space available to them due to habitat change counteracting
the positive (for these species) influence of climate change (Warren et al. 2001). Hence
whilst a warming climate may disproportionately affect some specialist species, it seems that
few species are able to fully utilise the potential opportunities for range expansion that it
creates.
4.1.3 Impacts of climate on reproduction and growth
Temperature affects insect development, survival and abundance. In temperate regions
temperature mainly acts by influencing winter survival, and increasing winter temperatures
may remove previous restrictions on a species’ establishment. For example, field transplants
involving the North American butterfly Atalopedes campestris, which has expanded its range
northwards, demonstrate that increases in winter temperatures have been particularly
important in its recent spread. Indeed, in some parts of its newly-colonised range summer
temperatures have remained relatively unchanged whilst winter temperatures have
increased (Crozier 2004). However, higher summer temperatures can also act by extending
the summer growth season for invertebrates (Bale et al. 2002).
4.2 Possible risks from assisted colonisation
Information on translocation risks for invertebrates is scant. Re-introductions have been
biased towards mammals and birds, and far fewer invertebrate re-introduction attempts have
been documented (Seddon et al. 2005). Most species that establish outside their native
range do so through accidental anthropogenic translocation.
Phytophagous invertebrates have the potential to become invasive when translocated to
areas where host plants have not evolved adequate defence mechanisms (Rhoades 1985).
Climate change may also lead to increased range sizes of pest species (e.g. Estay et al.
2009). However, as migrant insect species are expected to respond more rapidly to climate
change than plants (Lawton 1995), it may already be possible to assess whether this is likely
to be a substantial problem.
A low risk strategy (although not strictly AC) might be introductions into areas formerly
occupied by a species and which have become more suitable under a changing climate.
This may have particular relevance for those species that have previously retreated from
sites in the north of their range, possibly as a result of climate being only marginally suitable
and thus populations being more vulnerable to additional factors such as land use change.
Climate change may help to make populations translocated to these parts of their former
range more robust. Such an approach was proposed by (Carroll et al. 2009) who showed
23

that suitable climate space exists in the UK for the butterflies Black-veined White (Aporia
crataegi) and Mazarine Blue (Polyommatus semiargus) which both became extinct in the UK
in the early twentieth century. Under the JNCC (2003) guidelines this would be viewed as a
re-introduction, rather than as AC, as the species went extinct after the 1600 cut-off.
However with recipient sites being improved for the species through ameliorating climate
there are close obvious parallels with AC.
There is also likely to be lower risk of a species becoming invasive if its origin is
geographically close to the location of the recipient sites. In such cases the new site will be
likely to contain a similar suite of species to the source site, and new species interactions
with unpredictable consequences will be less likely to occur. This is analogous to reduced
invasive potential with reduced evolutionary distance between source and recipient
communities for vascular plants. An experiment on two UK butterfly species established
populations in suitable habitat 35 km north of their previous range. These species, Marbled
White (Melanargia galathea) and Small Skipper (Thymelicus sylvestris) have both been
extending their UK range northwards. However in each case, modeling suggested that
populations were not keeping track with climate and that the climate of the chosen recipient
sites was suitable for these species. In both cases the introduced populations survived and
grew over the first six year of monitoring. Furthermore, no detrimental effects on other
species in the recipient sites were detected (Willis et al. 2009), although the close
geographic distance between source and recipient communities would, in any case, make
this highly unlikely.
Negative ecological or economic impacts are usually only detectable for a small proportion of
alien terrestrial invertebrate species. For example, none was apparent among forty-two
established alien Heteroptera in Europe (Rabitsch 2008). However, at constant latitude, the
diversity of insect herbivores and the intensity of herbivory both increase with rising
temperatures (Bale et al. 2002). Hence, there is the potential for increased negative impacts
from both introduced and indigenous phytophagous species in a warming climate, and this
should be carefully considered in AC projects.
4.3 Factors determining the success of assisted colonisation
There are too few studies of invertebrate AC to develop any common rules in terms of what
regulates their success, and there appear to be few factors that commonly determine the
success of insect introductions in general. Clearly habitat suitability is crucial, and knowledge
of a species’ autecology is important in determining its habitat requirements (e.g. Marttila et
al. 1997). The extent of suitable habitat is also important. This is especially true for those
insect species which reproduce at high rates but which may be prone to population
fluctuations. To maximise the chance of success of translocations involving such species,
the release locality should contain a sufficiently large area of suitable habitat and the project
should involve a large founder population released during a climatically favourable period
(Hochkirch et al. 2007).
Even with apparently suitable habitat, it is crucial to determine the current climate suitability
of introduction sites as this can impact on the likelihood of introduction success. This may be
especially important in the UK, with many insect species reaching a northern range limit
within England or Scotland. Indeed, success of UK butterfly introduction projects has been
shown to be positively correlated with modelled current climate suitability (Menéndez et al.
2006) indicating that attempts to establish populations in sites that may become climatically
suitable, but which currently are not, may be inadvisable.
Insect (re)-introductions can be hampered by lack of fitness of stock used for re-introduction.
For example, some populations may be more able than others to withstand climatic
24

conditions that prevail at the edge of a range (Nicholls & Pullin 2000). Populations at the rear
(low altitude or latitude) edge of a range may respond rather differently to conservation
measures compared to those at the front of the range (Hampe & Petit 2005). To ensure
maximum likelihood of success, therefore, it has been suggested that insect releases should
ordinarily comprise individuals from a donor population that is located as close as possible to
the release site (Hochkirch et al. 2007). However, this may not account for releases as part
of an AC process. Furthermore, especially if dispersal is limited, some populations may have
become relics in gradually fragmenting habitat and, as a result, may have lost genetic
diversity (e.g. Ellis et al. 2006) whilst rear edge populations may remain long term stores of
greater genetic diversity (Hampe & Petit 2005). Under these situations, the geographically
closest populations may not be the fittest for use in AC and consideration might then be
given to carrying out introductions using potentially fitter populations from further afield. One
novel re-introduction to Britain, that of the recently extinct Short-haired Bumblebee (Bombus
subterraneus), is using queen bees from a population that had been introduced from the UK
to New Zealand in the late nineteenth century. Although clearly a source of stock with the
ideal genetic characteristics may not be available to many projects, it is hoped in this case
that the chances of successful re-establishment will be enhanced as a result (Gammans et
al. 2009).
Many introduced alien invertebrates are presumed to have become “permanently”
established. However information on long-term success of deliberately translocated
populations is rather scarce, especially for populations introduced to isolated habitat
patches. A notable exception is the butterfly Mountain Ringlet (Erebia epiphron) of the
endemic Czech race, E.e. silesiana. Two batches of fifty gravid females were released some
150 km from their natural range in 1932 and 1933. Whilst introduction at one site failed, at
the other the species became established and reached population densities similar to those
of the source population (Cizek et al. 2003). After seventy years, this introduced population
was found to contain most of the allozyme diversity of the source population. This suggests
that the initial introductions contained a sufficiently large number of founders with a sufficient
range of genetic variability to ensure fitness of the introduced population (Schmitt et al.
2005).
Most previous introductions of insects into sites outside of their range have been
unintentional. However translocation has been deliberately used to assist conservation of
several species of New Zealand wetas, a group of large Orthopteran insects. This has not
been for climate change related reasons, but rather because few or no sites within the native
range are free of introduced predators. Success has been mixed. For example Mahoenui
giant weta (Deinacrida mahoenui) were translocated to seven sites. Of the two
translocations that were successful, one was on an island with no previous records of the
species. In this case, the principal factor influencing establishment success was thought to
be absence of predation by rats (Watts & Thornburrow 2009). In a more extreme example, a
translocated population of Mercury Islands tusked weta (Motuweta isolata), which was
established on a previously unoccupied island, has now become the sole extant population
known after extinction of the species at the donor sight. Thus, this species was saved from
extinction by introduction outside the indigenous range (Stringer & Chappell 2008). Other
species of New Zealand wetas have similarly been established in predator-free refuges.
Whilst it is sometimes hard to determine whether the species had previously occurred in that
location (Watts et al. 2008) introductions outside the known range have significantly
improved the conservation status of several species. Similarly, in the case of the Czech
butterfly, E.e. silesiana (referred to above), the species occupies a small altitudinal range on
mountain tops at the donor sites. At the introduction site the mountains are 100 m higher
than at the “natural” sites, and thus the species may be buffered for a longer time against the
effects of a warming climate (Schmitt et al. 2005).
25

Of the few generalisations that can be drawn in terms of selecting candidates for AC, one
potentially interesting observation is that insects introduced to the British Isles by man
showed a negative correlation between body size and probability of establishment (contra to
introductions as a whole) (Lawton & Brown 1986). Supporting this observation, Heteroptera
that have established in Europe have a smaller body size than the European average
(Rabitsch 2008). However there is much variability within insect orders and the mechanisms
behind this observed pattern are not well understood. Hence whilst it might be considered
possible that smaller bodied species might be better candidates for successful AC, this
pattern is not likely to offer useful insight in individual cases.
4.4 Practicalities of translocating terrestrial invertebrates
The practicality of each insect introduction needs to be considered on its own merits.
However few reports of translocation projects targeting invertebrates have been formally
published. Bird and mammal cases accounted for 93% out of 180 animal re-introduction
projects identified from scientific literature by Fischer & Lindenmayer (2000). However some
aspects of AC projects involving invertebrates, for example animal housing costs, are likely
to be lower than those involving larger creatures. By way of example, a recent AC
experiment involved establishment of two butterflies in the UK. The modeling, release,
monitoring and follow-on analysis took around 8 person-months per species, of which about
half could have been undertaken by amateur volunteers. Additional project expenses were
less than £5000 (Willis et al. 2009). Furthermore, and as noted when discussing the possible
costs of applying AC, detailed knowledge of the current range, population trends and
autecology for a species are likely to be crucial for any conservation action including AC
(e.g. New & Sands 2004).
4.5 Summary & suggestions for experimental studies
As for the other species groups, our current information deficit for invertebrates is
substantial. Key uncertainties that might be addressed by experimental studies are:
Relationship of introduced target species to parasites: Some insects have complex
relationships with parasites, and indeed some host-specific parasites may play an
important role in regulating host populations. For example the parasitic
hymenopteran wasp, Cotesia bignellii, appears to be a specialist solely on the UKthreatened
Marsh Fritillary butterfly (Euphydryas aurinia), although there is debate
over the role that the species may play in controlling populations of its host (Barnett &
Warren 1995). However, in general it could be postulated that insect species with
specialist parasites may be more likely to become invasive if introduced outside their
current range in the absence of the parasite.
Environmental factors currently limiting range of focal species: Whilst we may
hypothesise that some species are able to benefit from being moved into newlyavailable
suitable climate space, other species, controlled by a more complex suite of
factors, may be unable to utilise the new location. For example the development of
some larvae in particular may be too closely tied to photoperiod to be able to adapt to
a new site if AC takes place over too great a latitudinal range (e.g. Lutz 1974).
Clearly, suitable food items need to exist for an invertebrate species to become
successfully established. Moving a species into habitat similar to that of its current
range should ensure the availability of appropriate food plants for phytophagous
invertebrates. However there may be interactions with plant quality (some species
may only be able to exploit a plant when it reaches a sufficient nutritional threshold;
26

e.g. Kerslake et al. 1998) and novel species combinations may result in new
competitive interactions. Furthermore new food plants may become available, with
the risk of promoting invasion by the introduced population.
Introduced populations may be especially vulnerable to predation, particularly in the
early phases of the project (e.g. Wagner et al. 2005). The risk of being preyed upon
by species that either do not occur or occur at different densities within the current
range and the AC location would need to be clarified.
Trial target species might be those with poor dispersal ability that have failed to keep up with
climate change. At the same time they might be rare enough to be of interest but not so rare
as to be difficult to work with, and where there would be substantial concerns about harm to
the source population. In all probability, this would mean species with at least moderately
specialised habitat requirements such that current populations are isolated from habitat
patches that may become suitable through climate change.
27

5 SPECIES DISTRIBUTION MODELS: PREDICTING SITES FOR ASSISTED
COLONISATION
5.1 Introduction
Climate change, land use change and their interactions may make species conservation
more challenging in the next few decades, mainly by driving changes in species
distributions.
There is evidence that distribution shifts are already occurring in directions consistent with
temperature increases recorded over the few last decades (e.g. Parmesan & Yohe 2003,
Root et al. 2003, Parmesan 2006) and that several species are moving towards the poles or
to higher elevation (Parmesan et al. 1999, Thomas & Lennon 1999, Walther et al. 2007,
Bergamini et al. 2009, Lättman et al. 2009). Despite this, and despite the need for sound
information on which to base conservation policy and planning, our ability to predict the
future is limited both by a lack of ecological data and an incomplete understanding of the
drivers of change.
Species distribution data are often not available or are incomplete, especially at local and
landscape scales. This problem can be partly alleviated by using models to predict habitat
suitability at locations that have not been surveyed (e.g. Martinez et al. 2006, Vanreusel et
al. 2007, Ohse et al. 2009). A plethora of models have been applied. Most of them can be
classified as empirical-correlative, process-based, or a combination of the two (e.g. Iverson
et al. 2004). The former are based on finding associations between variables believed to be
important and species occurrence or abundance (e.g. Guisan & Zimmerman 2000, Didier et
al. 2009, Tirpack et al. 2009). Such associations are either learnt statistically (Austin et al.
2009, Luoto et al. 2005) or by using machine learning methods such as Artificial Neural
Networks (e.g. Araujo et al. 2005) or genetic algorithms (e.g. Stockwell 2006, Termansen
2006 et al. Benito et al. 2009). Such models are used in isolation to understand drivers of
current distribution or coupled with scenarios of climate change to predict future distributional
change (e.g. Novaceck & Cleland 2001, Gomez-Aparicio et al. 2008, Dormann et al. 2008).
This can often be valuable when used in a comparative fashion, namely to assess
alternative policies and conservation strategies (Nelson et al. 2008).
Unlike empirical models, process-based models simulate biological events with degrees of
approximation that go from aggregated population models, relying on differential equations
(e.g. Kelpin et al. 2000), to spatially explicit individual based models simulating key events in
a species’ life-cycle, such as reaction-diffusion models and cellular automata approaches
(e.g. Crowley et al. 2005).
The relevance of this to the use of AC as a species conservation tool during climate change
is that discussion of climate-change driven AC commonly makes a number of assumptions
about the use of ecological modeling, and in particular the type of modeling described in
brief above that has been applied to predicting the response of species to climate change. It
is commonly assumed in the AC literature that we can successfully predict climatically
suitable current and future areas, central to assessing whether a species’ current range will
be completely unsuitable in the future, whether suitable climate space will exist elsewhere,
and where that climate space will be. Because of the importance of such assessments in
determining whether and how to undertake AC, it is essential that we base them on sound
evidence. In this section, therefore, we provide a thorough review of the range of possible
modeling techniques that might be applied to the issue of AC, enabling us to assess
whether, and under what circumstances, this reliance on modeling will be justified.
Specifically, we summarise some of the key modeling approaches that are used to attempt
to predict the responses of biodiversity to climate change, along with their associated pros
28

and cons, and discuss their application to the problem of determining the future possible
ranges of threatened species.
5.2 Species-habitat models
Species-habitat models (like climate envelope models, below) are built on the niche concept.
It can be argued that at equilibrium, a species’ distribution is constrained by resources and
community interactions such that these models estimate the ‘realised niche’ sensu
Hutchinson (1957). Their application does, of course, depend on this assumption holding.
Species-habitat models and climate envelope models differ only in the spatial scales over
which they are typically applied – there is no great conceptual or practical modeling
distinction. In general it is possible to see local, landscape scale and biogeographic scale
models as part of a hierarchy (e.g. Pearson et al. 2004), determined by (i) limiting factors
that act on eco-physiology (e.g. temperature, water, light availability), (ii) resources and (iii)
disturbance (including any type of perturbation of the other two), respectively (Guisan &
Zimmerman 2000, Morrison et al. 2006). Henceforward we refer to these factors as LRD (the
acronym of the variables (i) to (iii)). LRD are related to fitness, and determine ‘source’
habitats (Pulliam 1988) i.e. those areas or habitats in which a species is able to have a nondeclining
population. Different patterns emerge at different scales, with climatic variables
generally responsible for gradual changes over large areas, and landscape structure
responsible for distribution changes over smaller areas. Guisan and Thuiller (2005) offer a
helpful framework to reconcile different scales. In particular they point out that models
accounting for climatic factors alone can only deal – conceptually - with the potential range,
while they do not include any variables related to constraints operating at the landscape or
the local scale.
5.3 Climate envelope models
The usual models used to forecast potential distribution shifts at macro scales are commonly
referred to as ‘climate-envelope’ models (CEMs). These are built using the observed
relationships between present climatic variables and distribution data for the whole or part of
the species range. Modelled future climate data are then used as inputs to project potential
future species distributions, given the present relationships. Results from such models have
been used, for example, to predict the formation of new species assemblages at the
continental scale (Thuiller et al. 2005, Araujo et al. 2006), the number of species threatened
with climate-induced extinctions (Thomas et al. 2004), and to assess the robustness of
existing protected-area networks in view of climate change (Araujo et al. 2005, Hannah et al.
2007, Vos et al. 2008).
5.3.1 Modeling approaches
The simplest models are Generalised Linear Models (GLMs) (Nelder & Wedderburn 1972,
Thuiller et al, 2005) and Generalised Additive Models (GAMs) (Hastie & Tibshirani 1990,
Thuiller 2003, Platts et al. 2008). The first relies on presence/absence data, while the second
is a more flexible method that can also handle abundance.
Machine learning methods such as Artificial Neural Networks (ANN) (e.g. Pearson et al.
2002, Thuiller 2003) and Genetic Algorithms (GA) (e.g. Anderson et al. 2003, Termansen et
al. 2006, Fitzpatrick et al. 2007) are usually based on a training data set and a validation
data set. Classification–Regression Trees (CART) (Breiman et al. 1984, Bourg et al. 2005)
and Random Forest (which averages many CARTs) (e.g. Cutler et al. 2007) can be thought
of as convergence between machine learning and statistical methods, and are often used in
data mining in many applications. The same can be said for Bayesian Belief Networks
29

(McNay et al. 2006, Smith et al. 2007), that use Bayesian statistical theory to learn
relationships between the variables involved in a model.
The advantage of the above models is that they are usually straightforward to apply, and can
be used to forecast distribution shifts of a high number of species with little knowledge of
their life history. This can be an apparent strength in practical applications, given the lack of
resources and time to develop ecological understanding about hundreds or even thousands
of species. Such simplicity explains their widespread adoption by researchers and policy
makers. However, it can also turn out to be a weakness (see “Limitations” below).
Several studies have compared the performance of different modeling approaches (Olden &
Jackson 2002, Elith et al. 2006, Prasad et al. 2006). The results generally indicate that more
sophisticated models perform better (Elith et al. 2006) and that, ideally, the variables
included in a model should reflect knowledge about the species’ ecology (e.g. Guisan et al.
2006). This is often difficult in practice, and is linked to some of the limitations of empirical
models that we now discuss.
5.3.2 Limitations
Ecological and statistical theory is often not reflected well in these correlative statistical
models. While some studies exist regarding variable selection (Pearson et al. 2004), choice
of response curves (Austin & Gaywood 1994), or assessment of model assumptions in
making distribution projections (Araujo et al. 2005), although widely applied, these models
are nonetheless limited in what they can tell us.
An important point to note is that CEMs assume equilibrium between species distribution
and climate. In general, most statistical models, even at a more localised scale, assume
equilibrium with the variables of interest. This is convenient, especially for projections, but
not necessarily true. For example, a study by Svenning & Skov (2004) comparing potential
versus realized range for tree species, found that the post-glacial recolonisation of Europe
still appears limited by dispersal. Indeed, it can be argued that equilibrium is likely to be the
exception rather than the rule because species are continually trying to catch up with
changing environments. This may also be demonstrated by the problem of extinction debt;
as discussed in Section 2 some species are still undergoing range contractions because
there has not yet been sufficient time for negatively-impacting factors to cause the loss of all
affected extant populations.
Also, it is easy to over-interpret CEM results. They can only be interpreted, at best, as
indicating the potential distributions. There are, however, possible problems with the correct
identification of climatic limits. It can be difficult to secure an unbiased estimate of bioclimatic
limits in the presence of other unmeasured constraining factors operating at a local scale.
This is an example of the conceptual limitation of any empirical model, namely that it is not
possible to prove that correlation implies causation. Its practical consequence is that, (i) the
climatic variables might not be causally linked to the limits of an observed distribution and (ii)
if the climatic variables included in CEM are merely proxies for real causal factors, there is a
risk that future climatic change will not translate into the distributional changes predicted by
the model, unless the relationship between climatic proxies and putative causal variables
remains unchanged. Careful judgement is therefore needed to evaluate if the implicit
assumption that the observed species distribution is in equilibrium with climate is likely to
hold, at least as a reasonable approximation. As discussed in the sections examining AC
and the different species groups, this might not be the case when land use effects and
human intervention have greatly reduced a species’ range, in particular for some rare
species of conservation interest.
30

Furthermore, there are also statistical difficulties with CEM, many of which are beginning to
be addressed and others which so far have largely been ignored. We outline some of these
below.
5.3.3 Model selection and validation
Given that it is not known a priori which climatic variables limit a species’ distribution, there is
a risk of over-fitting the model by unwittingly including variables that have no real
explanatory power. Model selection is therefore a very important step. Researchers have
historically used two main strategies. Different, competing, models have been compared to
each other using an information theoretical framework or what is known commonly as the
Akaike Information Criterion (AIC) (Akaike 1974). Also, models have often been crossvalidated,
by using part of the data set as ‘training set’ and part as ‘validation set’ using
ROC/AUC (ROC: Receiver-Operator Curve; AUC: Area Under the Curve) as an evaluation
measure (e.g Luoto et al. 2005, Pearson et al. 2006, Wisz et al. 2008). This is the current
standard procedure but it is flawed, as we discuss below (see spatial autocorrelation).
Importantly, if model selection is applied (comparison of different models containing different
sets of covariates), the method adopted for model selection influences which variables are
included in a model, and hence predictions can therefore differ between models according to
selection method used. This problem has recently been addressed using ensemble models.
5.3.4 Ensemble models
Different methods perform differently from each other in predicting current (and therefore
future) species distributions (e.g. Elith et al. 2006, Lawler et al. 2006, Guisan et al. 2007,
Heikkinen et al. 2007) and habitat suitability (e.g., Evangelista et al. 2008, Roura-Pascual et
al. 2009). This is due to environmental variable selection methods (e.g Araujo & Guisan,
2006, Elith et al. 2006, Heikkinen et al. 2006, Dormann et al. 2008), sample size, and
correlations between variables involved (e.g. Guisan et al. 2006, Heikkinen et al. 2006,
Dormann et al. 2008, Graham et al. 2008).
Ensemble models have been introduced to alleviate such problems (e.g. Araujo & New
2007). A forecast ensemble consists of multiple simulations using different sets of initial
conditions (IC), model parameters (MP), and model classes (MC). Each combination of IC,
MC, and MP is one possible state of the system being forecasted. This method is relatively
new to distribution studies (e.g. Araujo et al. 2006, Pearson et al. 2006, Lawler et al. 2009,
Marmion et al. 2009). ANN, GARP, and Random Forest work using the ensemble principle
and select the best model out of many using some form of cross-validation. However, the
relatively recent novelty is in the combination of forecasts. Instead of choosing a ’best
model’, the range of projections from multiple models is explored and combined (e.g.
Crossman & Bass 2008, Barbet-Massin et al. 2009, Lawler et al. 2009).
Several studies quoted above have looked at the effect of using different methods to predict
species distributions, and some have devoted attention to comparisons between different
Global Circulation Models (e.g. Araujo et al. 2006, Broennimann et al. 2006, Mika et al.
2008, Barbet-Massin et al. 2009, Lawler et al. 2009).
Little has been done so far to quantify the variability between projections from ensembles
combining different sources of uncertainty (see Hartley et al. 2006, for an exception) and/or
to quantify the contribution of different sources to the total uncertainty (but see Dormann et
al. 2008).
5.3.5 Spatial autocorrelation
As mentioned above, a common tool to evaluate model-fit for presence-absence data is the
ROC (Receiver-Operator Curve). This is a plot of true positive cases vs false positive cases,
an assessment technique borrowed from signal processing theory. The standard and very
widespread use of the area under the curve (AUC) as a measure of model goodness of fit is
31

not without problems. Both statistical and machine-learning methods are not robust to nonindependent
errors, which can easily arise when the variables involved are spatially
autocorrelated. Although this topic is very well known and amply discussed in the spatial
ecology literature, most studies involving CEMs have ignored spatial autocorrelation as a
source of model error. This is worrying because for niche-based models significance can be
inflated up to a factor of 90 (Segurado et al. 2006). Beale et al (2008) examined this problem
in detail for European birds and were able to offer a methodological solution, based on
spatial randomisations of the observed distributions (see also Lennon 2000, Gimona &
Fernandes 2004, for similar early approaches). Other solutions, based on fitting models
explicitly including spatially correlated residuals, exist (e.g. Gimona & Brewer 2006, Beale et
al. 2007, Dormann 2007, Dormann et al. 2007, Hawkins et al. 2007, Bini et al. 2009). These
methods help select a model with fewer spurious relationships between distribution and
climate, and therefore in making more reliable predictions.
5.4 Landscape scale and local scale models
Landscape ecologists often partition space into ‘patches’ – more or less homogenous areas
of landscape – but it is possible to conceive of gradients (over landscapes) of the above
factors, defining ‘suitability surfaces’ (Mitchell et al. 2002, Carbonell et al. 2003, Fischer &
Lindenmayer 2006, Hossack et al. 2009, McGarigal et al. 2009).
Species-habitat models attempt to establish a formal link between some measures of LRD
and fitness. Such measures are usually defined at a spatial and temporal scales judged to
be appropriate and feasible. These models have been applied to mammals (e.g. Carrol et al.
1999), birds (e.g. Dettmers & Bart 1999, Graf et al. 2005, Kudo et al. 2005), insects (e.g.
Binzenhofer et al. 2005, Matern et al. 2007, Littlewood & Young 2008) and plants (e.g.
Guisan et al. 1998, Collingham et al. 2000, Peppler-Lisbach & Schroder 2004)
Often, abundance or species’ presence is used as a surrogate for reproductive output, and
the potential pitfalls of this pragmatic substitution have been pointed out for a long time (e.g.
Van Horne 1983, Pulliam 1988). Also, habitat models classify areas according to their
‘potential’ to be habitat, but misclassification error can be introduced by the fact that not all
suitable habitat is occupied (e.g. Pulliam 1996). In addition, as discussed with respect to
vascular plants in Section 3, habitat suitable for adults may not contain suitable
establishment conditions.
Scaling up is a challenge, as mapping LRDs is not always straightforward, especially if the
local resolution required is considerable (e.g. if there is complex micro-topography).
Integration of remote sensing and local measurements using modeling tools such as those
described above can be effective in building landscape-scale distribution models (e.g.
Nelson et al. 2005). Variables that can be measured remotely or are estimated from surveys
and interpolation can become part of models predicting distributions (e.g. Stohlgren et al.
1997, Poulin et al. 2002, Parviainen et al. 2009, Price et al. 2009). Modern sensors are able
to record variables such as vegetation structure, soil moisture and terrain characteristics with
resolutions that vary between centimetres and tens of metres. Laser-based terrain scanning
methods (LIDAR) are able to measure vegetation and terrain structure and are becoming
increasingly popular to model distributions of a number of taxa (e.g. Hill & Thomson 2005,
Sellars & Jolls 2007, Andrew & Ustin 2009, Muller et al. 2009, Seavy et al. 2009). However,
ground survey and validation efforts are important (e.g. Littlewood & Young 2008, Sesnie et
al. 2008) and often not trivial.
Finally it is worth mentioning that, when virtually no study is available on a target species, a
suitability index is sometimes calculated through ‘expert opinion’, for example by
conservation agency experts. Methods to elicit and combine knowledge have been proposed
32

(e.g. Store & Jokimaki 2003). This is often far from precise, but might give results that
compare well with model-based measures on sparsely sampled surveys, based on the ability
to predict presence and absence of a species. Johnson and Gillingham (2004) discuss
sources of uncertainty and point out ways to communicate it.
5.5 Mechanistic process-based models
CEMs provide at best a broad scale view of the potential future distribution range of species,
but do not include important factors such as how land use change is likely to influence
habitat availability or the effect of community/food web interactions, and dispersal (e.g.
Opdam & Wascher 2004, Hall et al. 2005, Travis et al. 2005, Butterfield 2009, Brooker et al.
2009, Gilg et al. 2009).
Theoretically, process-based models are certainly more robust in their description because
they can deal with biological processes and interactions, sometimes accounting for both
space and different trophic levels (Cassey et al. 2006, McCann et al. 2005, Gray et al. 2005,
Cagnolo 2009, Haynes et al. 2009). It could be argued that these should be a starting point
for modeling climate change effects.
Habitat suitability models and many community and food-web models concentrate on the
local characteristics shaping fitness and coexistence, and tend to disregard demographic
processes that happen at the landscape scale. Landscape scale process, such as
fragmentation and dispersal, however, can play an important role, as discussed in the
species group descriptions. Also, habitat fragmentation and climate change are likely to
interact, and, in this respect too, CEMs, assuming no dispersal limitation, and local models
assuming a closed population or community, can be insufficient to guide management.
Metapopulation models (e.g. Levins 1969, Hanski 1981) represent a class of mechanistic
models based on the balance between local colonization and local extinction rate (the
colonization rate of a site depends on the degree of occupancy in the surrounding
landscape, and on dispersal rate). These factors simulate the probability of extinction and
colonization events, which in turn are related to what some community ecologists call
‘propagule pressure’ (e.g. Tilman 1997). Various modeling studies have shown that there is
a threshold of patch occupancy below which species are committed to extinction (e.g.
Bascompte & Sole' 1996, Hill & Caswell 1999). An interesting early example of a study
looking at this issue is Collingham and Huntley (2000) who showed that when landscapescale
habitat availability for Tilia cordata falls below 25% its ability to migrate is curtailed
substantially. Travis (2003) showed that the impacts of habitat fragmentation may be
exacerbated by climate change, so the ability to keep pace with climatic changes is lessened
in the presence of habitat fragmentation and loss. The colonisation rate may be more
important than extinction rate when both climate change and habitat fragmentation are at
work. Perhaps unsurprisingly, species with a restricted range and poor dispersal abilities are
identified as more vulnerable to climate change (see also e.g. Opdam & Washer 2004,
Thomas et al. 2004, Best et al. 2007).
Colonising new areas does not mean simply arriving at (or being transported to) a site, but
also involves being able to persist and possibly grow. Local persistence is likely to depend
on the site conditions, determined in part by climate and in part by biological interactions. A
simulation study by Munkemuller et al. (2009) indicates that the relative sensitivity to climate
change and fragmentation can also depend on the mechanism of species coexistence.
Communities in which interactions are very weak persist better in moderately isolated
patches, and are less sensitive to climatic shift than fragmentation, as they rely on
recolonisation after random drifts to extinction. Communities in which coexistence is based
on component species having different mechanisms of regulation of density might be more
33

sensitive to range shift as their coexistence depends on large mean population densities to
avoid extinctions during density fluctuations.
Simple process-based mechanistic models have recently been combined with empirical
(CEM) models of range shift, to predict realised future distributions rather than their potential
extent. For example, Carroll (2007) used spatially explicit population models to investigate
likely land use and climate change effects on distributions. Iverson et al. (2008), incorporated
tree dispersal and establishment and forecasted distributional shift of a large number of
North American species. Schwartz et al. (2006) estimated extinction risk of trees and birds.
These authors used a combination of machine learning (Random Forest) and simple
mechanistic models, thus relaxing some of the CEM assumptions, such as no dispersal
limitation and no gradual variation in climate and landscape change. Incorporating such
processes is likely to be important when modeling at fine resolution (landscape level) but it is
not clear whether it is sufficient, as climate change is also likely to alter factors such as
population cycles, predation, and food web structure (Barton & Schmitz 2009).
Understanding details well enough to build biologically realistic models is a very expensive,
challenging and time-consuming task. Even simple mechanistic models can be difficult to
parameterise. When we consider more realistic models, which are necessarily parameterrich,
it is unlikely that sufficient data can be collected, certainly not across many species.
Accounting for competition, facilitation and multi-trophic interactions is clearly theoretically
sound, but at present seems most unfeasible for applications to most species. The basic
knowledge informing such models is often unavailable and decision making regarding policy
cannot wait several years. The role of complex models therefore seems more to explore
possible general management strategies in stylised scenarios (e.g. Hill et al. 2005, Vadadi-
Fulop et al. 2009) rather than to provide an accurate answer tailored to a particular case
study.
A combination of niche-based spatial models and metapopulation approaches might be the
only realistic (if very approximate) modeling option to make spatially-explicit predictions of
climatic impacts. For example, Keith et al. (2008) coupled a bioclimatic envelope and a
metapopulation approach to investigate the effects of various factors on the population
viability of South African fynbos plants. Anderson et al. (2009) used this approach on hare
species as a case study, and drew attention to modeling the trailing edge of the range during
climatic shift. They found that for relatively mobile species, higher dispersal capabilities delay
extinction at the trailing edge, probably due to a ‘rescue effect’ (sensu Hanski) from more
central patches. Populations at the core of a species range, in this model, therefore respond
differently to those at the margin, and those at the trailing margin of a shifting distribution
may lag relative to those at the centre. These authors also point out sources of uncertainty:
in population parameters, local adaptation, poorly known dispersal ability, small pockets of
ignored but climatically suitable habitat which may exist at finer resolution might yet be
important for management purposes.
5.5.1 Summary: the pros & cons of modeling approaches
In terms of Climate Envelope Models (CEMs), such as those used in the Monarch project
(Berry et al. 2005), the clear pros are that they are fast, explicit and (usually) simple to
understand. This provides policymakers and other stakeholders with a simple message (e.g.
a coloured map showing species range changes). In addition, many species can be
machine-processed quickly, thus making it easy to produce atlases of entire assemblages.
This simplicity, though, masks numerous serious problems.
CEMs are ecologically naïve: if they are not processed-based (i.e. the vast majority of
species distribution models) then there is usually no inclusion of ecological knowledge – in
fact, these models are typically pure data processing exercises. Furthermore, an association
is assumed between species distributions and environmental covariates (typically climate
34

and less commonly, land use). This is then ‘quantified’ (approximately) and then used to
predict distribution changes accompanying climate and land use change.
They are also statistically naive: CEMs as currently applied not only assume that climate
controls species distributions, but that it is the only factor of importance – this is clearly
inaccurate, as discussed in the species group sections (for example, we know that species
interactions matter, yet they are routinely ignored). The goodness of fit is generally AUC of
the ROC which is inadequate in that it does not test statistical significance of the climate
variables – the approach is flawed because AUC conflates spatial autocorrelation of the
variables with a mechanistic association. Finally, they also assume equilibrium between
species and environment, which may or may not be the case. It is certainly not the case
when important environmental covariates are changing. Overall, one of the major downsides
is that they give a veneer of reliability where little may be justified.
Process-based models (cf. invasion ecology modeling exercises) include more detail and
ecological realism on the focal species: this may be a much better guide to the feasibility of
modeling and predicting likely success of assisted translocation. These models are much
more likely to have a mechanistic component such that population dynamics - and perhaps
habitat associations based on species autecology - are included. At landscape and finer
scales there is much overlap between invasion models and metapopulation models.
Metapopulation models include realistic dispersal and colonisation, often with species
interactions (e.g. host plant dynamics coupled with herbivory).
5.6 Application of modeling approaches to assisted colonisation
5.6.1 Determining future responses to climate change
For the purposes of establishing the likely success of a translocation program a
metapopulation-like study of the focal species would be ideal: however the real world
constraints here are great, given the very large amount of fieldwork time and the necessity
for time-series observations of local extinction and population dynamics. Generally the
timescale for these studies would involve many generations of monitoring, obviously taking
many years.
Individuals surviving in their new environment and long term persistence might be difficult to
distinguish. There are numerous cases of individuals surviving outside their ‘normal’ range
(e.g. Carter & Prince 1981) but from a population perspective reproducing at less than
replacement (i.e. population λ

conceive of this phenomenon as one of populations going extinct or
starting, and data from one area show a big range expansion consisting of
new populations in 1977, followed for several years by range contraction
consisting of disappearance of populations.
What Simberloff does not emphasise here is that Carter & Prince (1981) found that plants
transplanted outside the range (northwards in UK) survived well, and that the ostensibly hard
edge to their range was an outcome of the interplay of local extinction and recolonisation –
that is, it is a population level phenomenon that may not be obvious from the behaviour of a
limited number of individuals.
In terms of AC, the combination of (i) the pros and cons of Species Distribution Modeling
(SDM), (ii) the very expensive time/effort needed for mechanistic metapopulation-like or
invasion-like modeling, (iii) the distinction between individual survival and population
survival, means that our options for making a robust case for the likely success of target
species must be couched within strong caveats. This is not to say nothing can be done. The
following may be an optimum combination of the pragmatic and the possible:
1. Statistically aware SDM. For example, appropriate techniques have been developed
jointly between MLURI and BioSS.
2. However, SDM sensu (1) are only possible if good data are available. Species will
have to be filtered on this basis.
3. Common garden experiments in several (preferable many and geographically widely
dispersed) locations would allow a space for time substitution.
4. Selecting the best locations to test the SDM - step (3) - could be informed by the SDM
itself.
5. Simple process based models may be possible given at the least some knowledge of
dispersal processes (although gaining knowledge of dispersal may be difficult).
5.6.2 Criteria for management-oriented modeling of species distribution
Given the review above, and general ecological considerations, when it comes to selecting
representative focal species or functional groups with the intent of modeling their fate for AC
management purposes, the information requirements are more detailed. For a ‘triage’ it
would be necessary to have information under as many as possible of the following
headings. Such information would help decide which groups should be focused on, and
which seem unlikely to be threatened, and would provide a basis for establishing research
needs.
Demographic and distribution traits
• Population abundance and its significance for extinction risk
• Recent population trends
• Factors likely to be limiting its present distribution
• The amount of fragmentation in the distribution
• Dispersal ability and propagule size
• Ability to recover from population decline (max growth rate)
Physiological traits
• Tolerance to projected changes in temperature
• Tolerance to projected changes in precipitation
• Phenology of the life cycle and vulnerabilities to CC
• Potential behavioural/in situ evolutionary adaptation to climate change
Habitat
• Habitats conditions needed for the full life cycle
36

• Information on which habitat components might be vulnerable to climate
change
• Presence of suitable habitat in the species projected range
• Likelihood that present habitat (e.g. vegetation type) can shift with climate
• Land use change impacts on habitat
A prioritization of species along these lines, to include a measure of uncertainty of our
confidence in the rank, would be highly desirable. Only then can the best returns on time and
money put into further modeling/analysis and subsequent informed management action be
secured.
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6 UNOCCUPIED HABITAT – AN ALTERNATIVE TO AC?
6.1 Are there empty niches?
An alternative to AC might be the translocation of individuals within the existing range but to
sites that are suitable (both now and in the future) and currently unoccupied. This would
negate concerns about moving the species outside of its range. Here we briefly consider
whether empty habitat (sometimes described as a “vacant niche”) might exist within a
species’ range, and whether it does indeed represent suitable recipient sites for
translocations to combat climate change.
Firstly, to clarify terminology, a species’ niche is the “multidimensional description of a
species’ resource needs, habitat requirements and environmental tolerances” (Hutchinson
1957). A species’ fundamental niche represents its absolute physiological tolerances to
abiotic factors such as temperature, nutrient levels, or water availability, and its realized
niche is the conditions under which it occurs when actually interacting with other species, i.e.
when a member of a community. A given species occupies a certain area in “niche space”.
When we suggest, therefore, that there is a “vacant niche” into which we can translocate
individuals, we are really discussing unoccupied suitable habitat – a particular niche is
already occupied by the target species, but it apparently fails to occupy all the habitat that
contains its niche.
Why then, might a species not occupy suitable habitat that occurs within its current range?
The composition of plant communities is seen classically as being regulated by a series of
filters which convert the global species pool into the realized plant community (Lortie et al.
2004). These filters are considered to work primarily at particular scales: biogeographical
events such as dispersal limitation determine the propagules that reach a site, abiotic
conditions regulate which species can survive, and local interactions determine which
species do survive.
However, all of these processes might also be operating at a very local scale to limit the
occupancy of suitable habitat. Firstly, abiotic and biotic conditions may fluctuate temporally
such that habitat which appears suitable at a given time of year is actually unsuitable at
other times, for example under seasonal grazing regimes. A clear case is the occasional but
lethal salt-water inundation of some recipient sites of the Sargent's cherry palm
Pseudophoenix sargenti (Maschinski & Duquesnel, 2007). In effect, these occasional events
mean that such habitats have actually been unsuitable overall, and will remain so unless
management can be imposed or the events are a stochastic one-off.
Even if the habitat is, overall, genuinely suitable, it may still be unoccupied. Firstly, dispersal
limitation can still be very important at the local scale. As discussed, dispersal limitation may
result from limited propagule production (possibly due to inbreeding) or habitat fragmentation
and a lack of suitable vectors (Honnay et al. 2002), or even simple factors such as dominant
wind directions being in the wrong direction. Furthermore, within metapopulations any given
population may be transient and dynamic such that suitable habitats will host repeated
cycles of establishment and extinction through time. Suitable but apparently-vacant habitat
may therefore simply be in the extinction phase of the cycle, and will ultimately be naturally
re-colonised at some future point. We should perhaps make a distinction, therefore, between
unoccupied and unutilized habitat – it is the latter that we are seeking – but we should still
not be surprised that unutilized habitat can exist within a species’ range given the common
impacts of dispersal limitation.
Finally if suitable unutilized habitat can be identified, might it be a suitable recipient location
for translocated individuals? Tautologically, if the habitat is currently suitable, then the
species should survive and reproduce there. The issue then becomes one of defining
38

whether that location will remain suitable in the future, and this returns us to the issues of
modeling future climatic suitability of a species’ current range, in particular in relation to data
limitation for rare species.
6.2 What is a species’ range?
Central to determining whether we are adopting an approach that utilizes unoccupied habitat
as opposed to AC is the distinction that the recipient site is either inside or outside of a
species’ range. Indeed, the definition of a species’ range is central to many of the arguments
concerning the practical application of AC. However, a clear definition of range is commonly
missing from discussion of AC, and of translocations in general.
An important distinction is that between the extent of occurrence (EoO) and the area of
occupancy (AoO). The IUCN Red List criteria guidelines (IUCN 2001) define EoO as:
“as the area contained within the shortest continuous imaginary boundary
which can be drawn to encompass all the known, inferred or projected sites
of present occurrence of a taxon, excluding cases of vagrancy…. This
measure may exclude discontinuities or disjunctions within the overall
distributions of taxa (e.g. large areas of obviously unsuitable habitat) [but see
'area of occupancy'… below]. Extent of occurrence can often be measured by
a minimum convex polygon (the smallest polygon in which no internal angle
exceeds 180 degrees and which contains all the sites of occurrence).”
In contrast AoO is defined as:
“…the area within its 'extent of occurrence'… which is occupied by a taxon,
excluding cases of vagrancy. The measure reflects the fact that a taxon will
not usually occur throughout the area of its extent of occurrence, which may
contain unsuitable or unoccupied habitats. In some cases (e.g. irreplaceable
colonial nesting sites, crucial feeding sites for migratory taxa) the area of
occupancy is the smallest area essential at any stage to the survival of
existing populations of a taxon. The size of the area of occupancy will be a
function of the scale at which it is measured, and should be at a scale
appropriate to relevant biological aspects of the taxon, the nature of threats
and the available data. To avoid inconsistencies and bias in assessments
caused by estimating area of occupancy at different scales, it may be
necessary to standardize estimates by applying a scale-correction factor. It is
difficult to give strict guidance on how standardization should be done
because different types of taxa have different scale-area relationships.”
As noted the AoO may not contain areas of suitable but unoccupied habitat. If the range of a
species is defined as AoO, then suitable unoccupied habitat might actually fall outside of the
species’ range.
This definition of AoO also highlights the question of scale. Whether or not a particular site is
within a species range may depend on the scale at which range is plotted. If the data is at
the 10 km square scale, large areas of land would be included in a plotted AoO that might
otherwise be excluded if the populations were mapped at a finer scale. For example, if we
undertook mapping at the scale of individuals on a particular mountain top - which for some
rare species would represent the full population of the species within Scotland - then we
would define a much smaller AoO, and would be more likely to exclude areas of suitable but
unoccupied habitat, even if they are a very short distance away. Certainly, given that the
39

guidance relating to undertaking translocations inside or outside of a range is quite different
(see section 7) there is a clear need to agree on the mapping scale at which range should be
determined in order to be able to appropriately classify any translocation activity.
40

7 UNDERSTANDING THE AC DEBATE, THE DECISION MAKING PROCESS,
AND THE POLICY CONTEXT
7.1 Understanding the AC debate and alternative decision-making approaches
At some point conservation agencies will be forced to make a decision on whether or not AC
should be implemented. Here we explore a number of recently-proposed frameworks for
both understanding the debate surrounding AC, and working through the decision making
process.
McLachlan et al. 2007 provide a multi-dimensional framework for debating AC between
involved parties, stating "It is important for academics, advocates, and managers to discuss
the role that assisted migration should play in the conservation of species." Their framework,
which may help those observing the debate to understand current stances, has three axes:
the perceived risk of AC, the perceived risk of inaction, and perceived confidence in
ecological understanding. Note that these axes are presented as independent but in reality
are unlikely to be so, i.e. improved ecological understanding will likely influence perceived
risk of both action and inaction. They then characterise three stances on this framework.
First, those aggressively favouring AC, second those against assisted migration, and third
those favouring “constrained assisted migration”, wherein AC is undertaken, but within a
science-led framework. It is possible, however, that most researchers would characterise
themselves within the third category, but would differ in terms of the direction and weighting
that they would give to the scientific evidence.
In a study of possible evolutionary responses to climate change, Skelly et al. (2007) discuss
a decision making process by which we might decide the risks to species from climate
change. From this we can deduce the following simple set of questions which might help us
determine whether AC is appropriate: Is there predicted to be an impact of climate change
on the species? If so, is there any possibility that the species may change its behaviour or
range? If not, is there any chance of evolutionary responses to enable persistence in situ? If
not, might AC then become a conservation option?
Hoegh-Gulberg et al. (2008) proposed a similar linear decision tree, where questions
concerning species threat and status are addressed sequentially. In their system, AC is only
one of the possible outcomes for managing species threatened by climate change.
Alternative outcomes include ex situ conservation (although it is always difficult to
comprehend at what point in the fluctuations of a dynamic environment we are likely to
consider re-introduction as being appropriate), and continuation of conventional conservation
action.
However, Richardson et al. (2009) criticise this approach, stating that it is simplistic to
believe that this kind of decision can be taken in a sequential manner. They argue that the
range of costs and benefits involved needs to be considered simultaneously. They propose a
multidimensional framework for evaluation, with four main categories of evaluation criteria:
focal impact (impact on focal unit and its community), collateral impact (impact on recipient
region), feasibility, and acceptability. The first three are all assessed on both ecological and
social criteria, and the fourth simply on the basis of social criteria. They say that this multicriteria
evaluation approach is an important improvement as "conservation decision-making
tools are most valuable when they help to distinguish the social and cultural values used to
judge acceptable risk from determination of risk itself".
It is hard to judge which of these approaches is most useful. Given that the categories of risk
and benefit are relatively similar in all cases, but that their magnitude differs between
species groups (as explored above) it seems sensible that each decision on whether or not
to proceed with AC should be made on a case-by-case basis (Vitt et al. 2009) but that the
41

assessment follows some form of standardised procedure or guidelines. It is also important
that any risk assessment or decision making procedures consider the risk of inactivity, and
weighs this against the risk of AC, as well as including different valuations of risks and
benefits by different stakeholders (Schlaepfer et al. 2009).
The conclusion of some contributors to the debate is that, no matter how much investment
we make in terms of understanding target systems etc., within acceptable time frames we
will not know enough to make the use of AC justifiable on the basis of objective risk
assessment. For example, Ricciardi & Simberloff (2009b) state that "All of these factors
[uncertainty of prediction and context-specificity of impacts] contribute to tremendous
uncertainty in the outcome of a species introduction and, thus, render risk assessments and
decision frameworks unreliable". In response it has been argued that "rejecting strategies
such as managed relocation based on the assertion that risk uncertainty is irreducible is
equivalent to putting one's head in the sand" (Schwartz et al. 2009). In addition it has also
been pointed out that it is likely to happen anyway under current legislative regimes, as in
the case of the Torreya Guardians (Fox 2007, McLachlan et al. 2009) and unregulated
butterfly translocations in the UK (Hodder & Bullock 1997). It is argued that, as restricting AC
is essentially unrealistic, it is better to have transparent discussion frameworks that can
account for multiple opinions (and differential weightings) when debating the pros and cons
of AC (Schlaepfer et al. 2009). However, in the section below we return to this issue of
“inevitability” of AC.
7.2 Conservation policy and guidelines relevant to assisted colonisation
Some conservation guidelines covering species translocations also discuss AC. At an
international level the most widely cited are those from the IUCN (1995) dealing with the
issue of re-introductions, following on from the IUCN (1987) position statement on the
translocation of living organisms (although there are some confusing changes in terminology
between the two documents – see Armstrong & Seddon 2008). The 1995 guidelines provide
a number of stages that should be worked through when planning a re-introduction. To
summarise, briefly, the key stages are:
Feasibility study and background research, including assessment of the species’
taxonomic status and an assessment of its critical needs, modeling of the release
populations, development, and habitat and population viability analyses.
Gather information and advice from previous re-introductions of the same/similar
species.
Select recipient site within the species’ historic range, with sufficient suitable habitat
and with reasons for previous decline or loss addressed. This may necessitate
habitat restoration prior to release.
Assess availability of suitable release stock. Individuals should only be removed from
wild populations after the effects have been assessed and guaranteed as nonharmful.
Socio-economic studies of the impacts to local human populations.
Re-introductions should comply with legislation of relevant country.
Design of pre and post release monitoring programmes including selection of
relevant success indicators, transport plans and release strategy.
With the exception of limiting release only to sites within the historic range, these are directly
applicable to AC. However, additional injunctions are recommended on the use of AC, in that
it should be considered “a feasible conservation tool only when there is no remaining area
left within a species’ historic range”, and that it “should be undertaken only as a last resort
when no opportunities for re-introduction into the original site or range exist and only when a
42

significant contribution of the conservation of the species will result”, although as we have
discussed (Section 6), definition of the current range may have its own problems.
In the UK the most relevant guidance comes from the 2003 JNCC document setting out a
policy for conservation translocations in Britain. The procedure for assessing the application
of translocations very much follows the IUCN 1995 guidelines. However, it also recommends
that AC be treated as the introduction of a non-native species, as detailed in Section 8.1.iii:
Species which are native to Great Britain (i.e. they are not believed to
have been introduced) and which still occur here, but where there is a
proposal to translocate individuals of the species beyond the current or
recent historic range (post 1600). This is actually an example of
translocation of a non-native species and is dealt with more fully by Anon.
(2003). An example would be the translocation of mammal species, such as
hedgehog, onto islands where they have not been recorded hitherto, such as
the Outer Hebrides (the consequences of a previous hedgehog translocation
to the Outer Hebrides have proved very damaging to populations of wading
birds, demonstrating the dangers of this category of translocation). Another
example would be to translocate buckthorn bushes to Scotland (this plant
currently only occurs as for north as Cumbria in northern England), for
example, with the intention of extending the range of an associated species
such as the brimstone butterfly. This category of translocation should be
subject to the full risk assessment procedure that applies to non-native
species, plus the use of the conservation evaluation procedure referred to at
8.1 (below). To receive approval, such proposals will need to meet the
requirements of each of these evaluations. This category of translocation may
become subject to additional legal controls in future to prevent damage being
caused to vulnerable species and habitats.
However, the JNCC (2003) guidelines also note (Section 7.2) that “translocation within Great
Britain of native species currently occurring here is not subject to significant legal controls or
constraints. However, translocation of wild species included on Schedules 5 and 8 of the
Wildlife and Countryside Act, 1981 and Schedules 2 and 4 of the Habitat Regulations, 1994
requires a licence from the relevant Country Agency.” These licenses are for the act of
“taking from the wild” rather than the act of translocation. As long as the species targeted for
AC is native and does not fall within one of these schedules (although many Scottish arcticalpine
species, for example, will be included), then it would appear that there is no legal
constraint within conservation legislation on such activity. It is hard then to assess by whom
approval might be given, other than by relevant land managers.
The JNCC report also states that, overall “The regulation of those activities which result in
the translocation of native species is generally best dealt with by codes of practice [such as
that from the IUCN, however…] …The exception to the preferred regulation of these
activities by such codes of practice relates to the translocation of species beyond their native
range. Because of the need to prevent damaging introductions to ecologically vulnerable
islands, stricter controls may need to be based upon new legislation to be effective in
preventing future problems. Such translocation of species beyond their native range is within
the scope of non-native species policy and is dealt with by the Defra Review of Non-native
Species Policy (Anon. 2003).”
The Defra Review (Anon. 2003) aims to “address the threats posed by invasive non-native
species without hindering legitimate activities”, although as noted above, the likely
invasiveness of an introduced species is hard to predict. Thorough and transparent risk
analyses are seen as essential, as well as the development of codes of conduct, which
should be given a “statutory underpinning”. This may, therefore, provide the statutory basis
43

for restricting AC, and preventing it from becoming “inevitable”. Although it might be argued
that AC species, particularly rare species, are not invasive and hence should not be covered
by such legislation, the current approach proposed in consultation on the Wildlife and
Natural Environment Bill (Scottish Government 2009) is that no non-native species can be
released into the wild unless they are explicitly exempted.
Another important point raised in the DEFRA report is that “Many problems posed by
invasive non-native species stem from a lack of public, commercial and institutional
understanding of the legislation prohibiting their release, and of the costs and consequences
of their establishment. Better information and education, and improved public awareness of
these issues are therefore all required. This should take account of translocation of native
species outside their natural range within Great Britain, which can also become invasive.”
This is closely related to the development of legislation to regulate AC. The need or desire
for such legislation appears to be driven by a fear of the consequences of unregulated and
unplanned AC, for example the release of hedgehogs on the Uists. However, actions such
as this, or those of the Torreya Guardians, might have been averted if there was a greater
public understanding of the risks to in situ species and habitats.
To summarise, although there are detailed guidelines, and although some target species or
associated donor or recipient habitats may have some form of legislative protection, there
are no specific legal restrictions on the within-country translocation of species in order to
undertake AC in the UK. This does not appear to be a situation which is unique to the UK
(McLachlan et al. 2007), reflecting the fact that conservation legislation (and associated
assessment of native status) commonly operates at a national level (Vitt et al. 2009).
However current consultation on new conservation legislation such as the Wildlife and
Natural Environment Bill indicates that this may soon change.
Given that such emphasis is placed on, ultimately, the IUCN guidelines, it is worth
considering the feasibility of their application. If they were strictly followed, the level of
information needed, both prior to and following translocation, would mean that even if the
data are available, any translocation project would be highly complex and costly. As
highlighted in Section 5, the issue of data availability is also not a trivial one. Assessing, for
example, the native status of a species or the inevitability of its demise in its current location
may be extremely difficult tasks. In addition, waiting until translocation has become a tool of
last resort may severely restrict its likely success, and reviews of the re-introduction literature
would suggest that following the guidelines does not necessarily help in terms of making the
translocation a success (Dalrymple et al. in Prep). Furthermore, although it has been
suggested that the application of these guidelines has generally been pragmatic, for
example in the case of translocations already undertaken in Scotland (C. Sydes pers.
comm.), it is questionable whether such pragmatism will still be applied when dealing with
AC.
Here, though, we should perhaps recognise that as long as there is no legal framework
preventing AC, and as long as the guidelines pertaining to AC appear to be difficult to
implement and not necessarily beneficial, the chances of AC acts that fail to follow such
guidelines will be increased.
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8 SUMMARY & SYNTHESIS
It is notable that the topics raised in the subsections examining the different species groups
closely match those discussed in the general introduction – to this end the problems of
applying AC in a Scottish context do not appear to be particularly unique.
At a first glance, we might readily state that candidate species for AC will have:
Certainty of extinction in current range
Fragmented distributions and/or limited dispersal potential
Suitable future climate space in Scotland
Limited capacity for in situ adaptation
However, although it is easy to list such attributes, applying them to lists of species is difficult
because:
In those cases where there is a strong link between climate and distribution, i.e.
mountain-top specialists, there may be a reduced chance of future suitable climate
space occurring in Scotland.
In those cases where suitable future climate space may exist, significant impacts of
other drivers of species decline, particularly habitat specialism and fragmentation,
limit the availability of data for modeling the link between distributions and climate.
Although generic rules concerning the links between species traits (e.g. seed size,
habitat specialism) and dispersal ability may hold true in some cases, in others these
relationships break down. Trait-based assessments may not, therefore, be fool-proof.
Likelihood of in situ adaptation at a sufficient rate is extremely hard to predict. It will,
however, certainly depend on the level of genetic diversity present and the potential
for its recombination. Genetic diversity is likely to be low, and gene flow limited, in
rare or infrequent species with limited distributions.
As well as risks from climate change, in combination with other factors, species might be
selected on the basis of characteristics that make the application of AC both acceptable and
likely to succeed. These include:
AC being the “last resort” for the species
Low impact on donor populations
Selection of sufficiently diverse and suitable genetic stock, and propagules of
particular types (e.g. vegetative).
Selection of appropriate site for introduction including suitable metapopulation
dynamics
Recreation of necessary biotic interactions
Low impact on recipient communities
Again, though, although such characteristics are easy to list, their application is complex
because:
We have very limited data on the long-term population-level trends of many of the
rare species that might be suitable candidates for AC, as well as populationregulating
characteristics and demographic processes.
We have no clear understanding of what constitutes the point of “last resort”. Should
it be the point at which the species is clearly on a trajectory to extinction, or should it
be the last point at which AC can be successfully undertaken? The former may be
hard to predict, but might already have been reached for species carrying some form
of extinction debt. The latter may involve much larger population sizes than we
currently suspect, especially given the other demands for sufficient genetic diversity
and limited impacts on the source populations.
Our lack of understanding of population dynamics makes it difficult to determine a
priori that there will be no impact on donor populations.
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We very rarely have any information on the genetic diversity of rare plant or animal
species. Even when we do, the amount of genetic diversity needed to create de novo
a healthily reproducing (and potentially adaptable) population is almost impossible to
predict.
We do not have an understanding of the best locations from within a current range
from which to select material for translocation.
We often do not understand the full suite of factors regulating population success,
and so cannot be sure that a site which appears to be suitable contains all of the
essential characteristics and does not contain inhibitory factors.
Our limited understanding of dispersal distances and genetic structure also limit our
ability to understand the occurrence of/need for meta-population dynamics.
Some species need very specific biotic interactions. These may be difficult to
recreate.
Although, for the species groups considered here within the Scottish context (either
rare or mountain-top species), the risks to recipient systems are likely to be relatively
low, any impacts are often unpredictable.
The central over-arching problem with respect to the application of AC is the possible
demand for a high level of knowledge and certainty, and the inherent lack of data for the vast
majority of the species that might be considered candidates. Although for (non-AC)
conservation translocations IUCN guidelines may be applied pragmatically, such pragmatism
might not extend to AC. Furthermore the data deficiencies that exist tally closely with the
data necessary for the sensible application of many of the possible modeling approaches
described in Section 5, as well as being in general the data necessary for any conservation
programme.
8.1 Areas for priority action
Although there is no shortage of possible targets for research projects exploring the
application of AC, we would suggest that the following are priority areas for action:
Developing assessment processes to attempt to identify candidates for AC: possibly
using a risk assessment process that combines trait data (e.g. for dispersal ability,
reproductive characters, growth form) to assess likely impacts of climate change and
habitat fragmentation - along with risk of invasiveness - with modeling data on likely
occurrence of suitable future climate space and/or potential risk of decreasing
landscape connectivity from further habitat fragmentation. Data required may include
that listed under the proposal for a form of species triage (Section 5).
Investigations of dispersal abilities and genetics, for example in arctic alpine lichen
species, would be very useful to determine if certain groups of species would benefit
from assisted colonisation.
Practical trials of transplant methodologies for particular species groups, e.g.
terricolous and saxicolous lichen species. These should probably focus on use of
propagules rather than adult thalli if possible. Any field trials should be carried out
with properly replicated, statistically sound methodologies and long term monitoring
of both donor and recipient systems.
Better understanding of what constitutes “last resort”, and how this is influenced by
e.g. genetic diversity and breeding systems.
Exploring the best locations from which to source material for AC relative to current
and future climatic conditions.
Combining climate projections and knowledge of vegetation transitions to select sites
that will be suitable for future survival of the species. Should we actually select
recipient sites that look like the current habitat of the species?
46

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