Download Evolutionary responses to global change: lessons from invasive

Ecology Letters, (2014) 17: 637–649
doi: 10.1111/ele.12262
REVIEW AND
SYNTHESIS
Emily V. Moran,* and Jake M.
Alexander
ETH Zurich, Universitatstrasse 16,
8092, Zurich,Switzerland
*Correspondence:
E-mail: [email protected]
Evolutionary responses to global change: lessons from invasive
species
Abstract
Biologists have recently devoted increasing attention to the role of rapid evolution in species’
responses to environmental change. However, it is still unclear what evolutionary responses should
be expected, at what rates, and whether evolution will save populations at risk of extinction. The
potential of biological invasions to provide useful insights has barely been realised, despite the
close analogies to species responding to global change, particularly climate change; in both cases,
populations encounter novel climatic and biotic selection pressures, with expected evolutionary
responses occurring over similar timescales. However, the analogy is not perfect, and invasive
species are perhaps best used as an upper bound on expected change. In this article, we review
what invasive species can and cannot teach us about likely evolutionary responses to global
change and the constraints on those responses. We also discuss the limitations of invasive species
as a model and outline directions for future research.
Keywords
Biotic interactions, climate change, cline, genetic constraints, invasive species, local adaptation,
niche, range expansion, rapid evolution, selection.
Ecology Letters (2014) 17: 637–649
INTRODUCTION
Over the past decade, biologists have devoted increasing
attention to the role of evolution in species’ responses to environmental change. Much of this interest has arisen from concerns about the ability of species to respond to rapid
anthropogenic global change (Davis et al. 2005; Jump &
Penuelas 2005; Bradshaw & Holzapfel 2006; Parmesan 2006;
Reusch & Wood 2007; Hendry et al. 2008). Indeed, human
influences have given rise to many classic examples of rapid
evolution, including industrial melanism, tolerance of heavy
metals or air pollutants and insecticide and herbivore resistance (reviewed by Reznick & Ghalambor 2001). Other studies have documented the rapid evolution of invasive species in
response to novel biotic and abiotic conditions, and of native
species in response to the invaders (Callaway & Maron 2006).
Furthermore, theoretical studies and evolutionary experiments
suggest that local adaptation can affect the ability of populations to grow, persist and colonise new habitat (Lavergne
et al. 2010).
These and other lines of evidence show that evolution may
be an important component of species’ responses to diverse
agents of global change (Visser 2008), especially given that the
pace of change is likely to limit species’ ecological responses,
such as migration to areas that are becoming climatically suitable following rapid climate change (Jump & Penuelas 2005).
Nevertheless, until recently both models and empirical studies
of global change responses largely ignored evolution
(Lavergne et al. 2010), and it remains a challenge to predict
what evolutionary responses can be expected and at what
rates they will occur. This is partly because of the difficulty of
performing experiments over relevant timescales and partly
because some of the most drastic challenges to species, such
as global climate change, are ongoing and are themselves difficult to predict. In this article, we consider the extent to which
the natural experiments offered by the spread of invasive species can help overcome these challenges and inform us about
likely evolutionary responses to environmental change in
native taxa. We focus particularly on climate change because
of the close parallels between the novel conditions experienced
by species introduced to new geographic areas and those
resulting from climate change, but many insights are applicable to other drivers of global changes.
Evolutionary changes consistent with responses to current
climate change have already been detected in some species,
including modest shifts towards earlier breeding phenology, in
critical photoperiod for diapause timing, in clines of climaterelated chromosomal inversions and in the proportion of
high-dispersal phenotypes in species undergoing range shifts
or expansions (Balanya et al. 2006; Bradshaw & Holzapfel
2006; Hanski 2012). These examples are as yet limited, in part
because distinguishing evolutionary change from plastic or
ecological responses is challenging (Skelly & Freidenburg
2010). Moreover, the degree of climate change experienced in
most areas of the world is still small compared to future
projections (IPCC 2007), so selection pressures may still be
too moderate to provoke strong evolutionary responses in
many species. Indeed, studies of several mammal and bird
species, most with generation times longer than a year, have
confirmed plastic phenotypic responses but not genetic
changes (Merila 2012); to date, genetic responses have mostly
been observed in species with generation times of 1 year or
© 2014 John Wiley & Sons Ltd/CNRS
638 Emily V. Moran and Jake M. Alexander
less (Reusch & Wood 2007; Merila 2012). Virtually all of
these studies have been based on longitudinal field observations – only rarely, as in the case of pitcher-plant mosquito
photoperiod responses, have the genetic change and its fitness
consequences been confirmed in the lab (Reusch & Wood
2007). Experimentation is inherently difficult, because evolutionary change is slow relative to the length of most experiments, and chronosequences (e.g. of seeds or spores) from
populations that have experienced climate change are not
available for most species. Thus, much controversy remains
regarding the extent to which evolutionary processes must be
considered when attempting to predict species and community
responses over the next 50–200 years (Davis et al. 2005;
Parmesan 2006; Lavergne et al. 2010).
Given these limitations, complementary approaches must be
considered to study evolutionary responses of species to climate change. One possibility is to exploit the large-scale
experiments provided by the transcontinental transport of
non-native species (Callaway & Maron 2006; Gilchrist & Lee
2007; Sax et al. 2007). A recent review examined what invasive species – i.e. species that are actively spreading in a new
region (Daehler 2001) – may teach us about the possibility
for, and consequences of, range shifts in response to climate
change (Caplat et al. 2013). As yet, however, limited attention
has been given to what invasive species may teach us about
evolutionary adaptation to climate change.
The parallels between the invasion of new environments
and the environmental changes associated with climate change
are strong, though imperfect. Many of the novel conditions
encountered by invasive species, including new temperature or
precipitation regimes, altered mutualistic or antagonistic biotic
interactions and the availability of new suitable habitat unoccupied by conspecifics, are similar to those that species will
confront following climate change (Table 1). Some of these
conditions apply to other global changes as well. Invasive
species and native species exposed to rapid environmental
change can also experience similar demographic perturbations.
For example, both newly founded non-native populations and
native populations in fragmented landscapes can experience
severe population bottlenecks, and range expansions of nonnative populations are analogous to expansions of native populations following climatic or land-use changes. In both cases,
comparative studies using invasive species could provide
insights into the role of gene flow, genetic variation and life
history traits for population recovery and adaptive evolution
following these demographic events. Although these processes
are also studied in the native range, one major advantage of
invasions is that they provide the opportunity to directly compare populations that have experienced changing environmental conditions during spread with those in the native range
that have not experienced those changes.
The ‘experiments’ offered by biotic invasions have other
advantages. Rapid evolution and eco-evolutionary interactions
frequently occur both in the invader and in the native community (Lambrinos 2004; Callaway & Maron 2006; Lankau
2012) and they do so in the context of multiple interacting
factors, rather than the simple single-factor conditions used in
most selection experiments (Reusch & Wood 2007). Moreover, biological invasions span the same timescales, from dec© 2014 John Wiley & Sons Ltd/CNRS
Review and Synthesis
ades to centuries, over which we attempt to predict global
change responses.
Before attempting to extrapolate from such data sets, however, one must keep in mind that successful invasive species
are, by definition, good at spreading and exploiting new environments. Not all introduced species become invasive, and
there are potentially important differences in the genetic structure of native and invasive populations and in the selective
pressures that they face; thus, many native species may exhibit
more constrained ecological or evolutionary responses than
do successful invaders. Nevertheless, invasive species may provide a useful upper bound to the amount of evolutionary
change we should expect to see in response to climate change.
We do not intend here to present an exhaustive review of
all studies of evolutionary change in invasive species, nor of
the theoretical background behind eco-evolutionary interactions in the context of climate change. These topics have been
reviewed previously, at least in part (Holt 1990; Lambrinos
2004; Huey et al. 2005; Callaway & Maron 2006; Reusch &
Wood 2007; Sax et al. 2007; Prentis et al. 2008; Visser 2008;
Lavergne et al. 2010; Matesanz et al. 2010). Rather, our aim
is to focus on how knowledge derived from species introductions can shed light on two major questions about evolutionary responses to climate change. First, which traits are likely
to exhibit evolutionary responses to the novel environments
encountered during invasion and rapid climate change, and
how rapidly? Second, what limits evolutionary responses to
novel environments, and is evolution likely to rescue populations at threat of extinction from climate change? We also discuss the potential limitations of using invasive species to
derive lessons applicable to native species experiencing environmental change and outline promising avenues for future
research.
WHAT KIND OF EVOLUTIONARY RESPONSES TO
CLIMATE CHANGE CAN WE EXPECT, AND HOW
QUICKLY?
Climate change is likely to alter selection pressures in several
ways. Many species exhibit local adaptation to climate, and as
a result frequently exhibit clinal distributions of climaterelated traits, such as size and reproductive timing, along climatic gradients. Changes in temperature and precipitation
regimes will select for shifts in these clines – for example, with
warm-climate phenotypes being favoured in historically colder
regions (Visser 2008). Even populations that are able to
migrate to areas with climates similar to their current habitat
will likely experience novel selection pressures due to altered
biotic interactions (Parmesan 2006; Williams & Jackson 2007),
as well as novel combinations of photoperiod cues and climate, which means that plastic responses alone are unlikely to
be sufficient in the long term (Visser 2008). In addition, while
the availability of newly suitable habitat beyond the current
range edge could select for increased dispersal ability (Phillips
et al. 2008; Hanski 2012), habitat fragmentation leading to
high dispersal costs may create an opposing selective force for
reduced dispersal in some populations (Lambrinos 2004;
Cheptou et al. 2008). Changes in average dispersal ability,
whether positive or negative, would in turn affect migration
Review and Synthesis
Invasion and evolution to global change 639
Table 1 Biological invasions as a model system for evolutionary responses
to selection pressures imposed by climate change. Examples can be found
in the text
Selection
pressure
Predicted response
in native species
Observations during
biological invasions
Novel climatic
conditions
Shifts in climate-related
traits to match new
local optima
Evolution of clinal trait
variation along climate
gradients (Table 2)
Availability of
habitat unoccupied
by conspecifics
Evolution of increased
dispersal ability to
exploit habitat made
suitable by climate
change
Evolution of increased
dispersal ability at
invasion front
Novel biotic
interactions
Adaptation to novel
antagonists/mutualists
encountered due to
range shifts
Evolution of native
species in response to
invasive species acting
as a novel host/
predator/prey/competitor
(Table 3)
ability, meta-population dynamics and gene flow, and thus
feed back to affect ecological and evolutionary responses to
climate. Invasive species have provided classic examples of
rapid evolution in traits related to climate, biotic interactions
and dispersal, and how quickly this evolution can occur.
Abiotic selection pressures: climate
One of the most commonly documented forms of rapid evolution in invasive species is the evolution of clines in climaterelated traits (Table 2). Invasive species can face strong selection pressures to match the local climate, especially when the
founders come from an area with a very different climate
(Alexander 2013). Because species differ in their life history,
the most useful measure of time is generations rather than
years. Examples from the invasion literature show that geographic clines generally develop 50–150 generations after
introduction, and occasionally in < 25 generations (Table 2).
Extensive genetic change can sometimes occur despite both
limited environmental variation in the native range or reduced
genetic diversity in the non-native range. For example, the
perennial shrub Hypericum canariense exhibits half as much
allelic diversity and a third of the heterozygosity in California
and Hawaii compared to native Canary Islands populations.
Despite this, invasive populations evolved a latitudinal cline
in flowering time that exceeds the variation in flowering time
seen in native populations and also exhibit an increase in
growth rate (Dlugosch & Parker 2008b).
The observed phenotypic changes documented for the examples in Table 2 were confirmed to be due to genetic (and/or
epigenetic) rather than plastic responses to different environments through common garden experiments. However, unless
multiple generations are studied (e.g. Huey et al. 2000), common garden experiments alone cannot distinguish between
genetic and epigenetic effects. More multi-generational common garden studies would therefore be desirable.
Although all of the genetic changes documented in Table 2
are consistent with local adaptation (Weber & Schmid 1998;
Lounibos et al. 2003; Dlugosch & Parker 2008b), random
drift or founder effects could also contribute to apparent
clines. Two methods used to test for local adaptation are reciprocal transplants to examine survival or reproduction in different environments (Quinn et al. 2001), or the comparison of
phenotypic clines to neutral expectations based on colonisation history (Keller et al. 2009). For Echinochloa crus-galli,
Hakam & Simon (2000) confirmed that the greater cold tolerance of northern populations was due to the higher activity of
a protective enzyme, and Roy et al. (2000) showed that these
populations had descended from southern cold-sensitive
North American populations. Either the direct measurement
of fitness or the use of null models should be applied more
widely when assessing evolutionary responses to invasion or
global change. Nevertheless, it should be possible to confirm,
in some cases from existing data, whether the parallel clines
frequently observed in both native and introduced ranges are
indeed adaptive. If so, further experiments, as in the Echinochloa example, could shed light on the mechanisms of adaptations. This in turn may help us identify what traits to focus
on when measuring genetic diversity in native populations
experiencing global change.
Abiotic selection pressures: habitat availability
For organisms encountering a large area of suitable but unoccupied habitat, high dispersal rates can be strongly advantageous (Fig. 1). If this is a widespread response in native
species, it could enhance species’ ability to track climate
Table 2 Examples of rapid evolution in climate-related traits in invasive species
Species
Common name
Trait
Years
Generations
Reference
Drosophila subobscura
Aedes albopictus (Fig. 1)
Drosophila subobscura
Hypericum canariense
Oncorhynchus tshawytscha
Solidago altissima
Eschscholzia californica
Lythrum salicaria (Fig. 1)
Silene vulgaris, S. latifolia
Solidago gigantea, S. canadensis
Echinochloa crus-galli
Fruit fly
Tiger mosquito
Fruit fly
Canary Isl. St. John’s wort
Chinook salmon
Late goldenrod
California poppy
Purple loosestrife
Campion
Goldenrod
Barnyard grass
Chromosomal inversions
Photoperiodic diapause
Wing size
Flowering phenology
Growth and reproductive traits
Growth traits
Flowering and growth traits
Time of and size at flowering
Various growth, flowering traits
Flowering time, growth traits
Photosynthetic enzyme activity
10–15
15
c. 20
<50
c. 90
c. 100
110–150
c. 200
c. 200
c. 250
<400
c. 50–150
25–150
c. 100
<25
<30
c. 20–60
110–150
c. 100–200
c. 100–200
c. 50–150
<400
Balanya et al. (2006)
Lounibos et al. (2003)
Huey et al. (2000)
Dlugosch & Parker (2008b)
Quinn et al. (2001)
Etterson et al. (2008)
Leger & Rice (2007)
Montague et al. (2008)
Keller et al. (2009)
Weber & Schmid (1998)
Hakam & Simon (2000);
Roy et al. (2000)
© 2014 John Wiley & Sons Ltd/CNRS
640 Emily V. Moran and Jake M. Alexander
Figure 1 Global change factors are expected to favour
dispersal ability (orange). Similar selection pressures
petiolata), the development of clines in climate-related
Photographs obtained from: Emily Moran (Rhinella),
(Aedes).
Review and Synthesis
adaptation to novel biotic interactions (green), shifts in climate-related traits (pink), and changes in
in the invasive range have led to changes in the production of allelopathic chemicals (Alliaria
traits (Lythrum salicaria & Aedes albopictus) and greater dispersal ability (Rhinella (Bufo) marinus).
Jake Alexander (Alliaria), Robert Coulautti (Lythrum) and the CDC Public Health Image Library
change. Other global changes could also lead to suitable habitat becoming available beyond the historical species range,
and therefore selection for increased dispersal, including nitrogen deposition (for N-limited species), deforestation or habitat
disruption (for ruderal or shade-intolerant species), relaxation
of hunting or other human activities or range expansion of
host species (due to a variety of global change factors). There
is much theoretical support for the evolution of increased dispersal ability at expanding range edges (Phillips et al. 2008),
but empirical evidence is still somewhat limited. A classic
example is the invasive cane toad in Australia (Fig. 1), which
at the range front has evolved longer legs and the ability to
travel longer distances (Phillips et al. 2008). An increase in
dispersal ability has also been documented in multiple insect
species that are currently expanding their native ranges due to
the wider availability of suitable habitat (Hanski 2012).
However, interactions with habitat fragmentation may hamper
the evolution of increased dispersal, as we will discuss later.
Invasive species present an opportunity of testing whether the
evolution of increased dispersal ability during range expansion
is a common phenomenon because of the wide range of invasive taxa and stages of invasion available for study. Similarly,
they provide opportunities for testing how long such increases
may persist after local densities increase – which is likely to
select for competitive ability.
Biotic selection pressures: novel species interactions
Changes in species interactions resulting from asynchronous
range shifts in response to climate change could induce selective pressures similar to those encountered during invasion, as
© 2014 John Wiley & Sons Ltd/CNRS
species are exposed to new predators, competitors and mutualists. Other agents of global change that cause the reduction
or loss of an interacting species (e.g. due to harvesting) and
alterations in competitive strength (e.g. due to differential
responses to N deposition) can also alter biotic interactions.
By studying the evolutionary consequences of novel interactions during invasions, we may gain important insights into
what traits or environmental factors promote adaptation (as
in the examples in Table 3) vs. lack of adaptation or exclusion. Many species do fail to adapt. For instance, while
Sporobolus airoides showed evidence of adaptation to Acroptilon repens allelopathy, Heterostipa comata (a similar native
grass) did not (Mealor & Hild 2007).
Evolutionary responses to biotic interactions can be as fast
as or faster than adaptation to local climate. Mortality caused
by an invader can be high, and this strong selection pressure
can lead to rapid changes – provided the population does not
go extinct first. Unsurprisingly, insects exhibit particularly
rapid responses: numerous species have not only evolved the
ability to feed on invasive plant or animal hosts, but also
show signs of reproductive isolation from populations feeding
on their original host (Table 3); adaptation to new hosts can
occur in < 40 generations. For instance, weevils (Euhrychiopsis
lecontei) found on invasive Eurasian water milfoil (Myriophyllum spicatum) show higher oviposition fidelity to this host
than do those found on the native water milfoil (M. sibericum), and laboratory crosses confirm that this preference is
genetically based and that fecundity is reduced in betweenhost crosses (Sheldon & Jones 2001). Native species have also
exhibited rapid evolution in response to the introduction of
new predators, unsuitable prey items or competitors within
Review and Synthesis
Invasion and evolution to global change 641
Table 3 Examples of rapid evolution in response to novel species interactions
Species
Common name
Host switching to invasive species
Cotesia glomerata
Parasitic wasp
Euhrychiopsis lecontei
Milfoil weevil
Euphydryas editha
Edith’s checkerspot
Jadera haematoloma
Soapberry bug
Rhagoletis pomonella
Apple maggot
Response to invasive predator or (unsuitable) prey
Mytilus edulis (Fig. 1)
Blue mussel
Trait
Years
Generations
Reference
Host use
Host use, fecundity on host
Host preference
Host use, mouthparts,
development rate
Host use, phenology
<10
<11
<20
<40
c. 30
c. 33
<20
c. 100
Tanaka et al. (2007)
Sheldon & Jones (2001)
Singer et al. (1993)
Carroll et al. (2001)
<150
<150
Filchak et al. (2000)
c. 15
<15
Freeman & Byers (2006)
<60
<60
Rana aurora
Northern redlegged frog
Induced shell thickening in
presence of predator
Hiding in presence of predator
Pseudechis porphyriacus
Red bellied black snake
Toxin tolerance, lower attack rate
50–67
c. 13–23
Kiesecker &
Blaustein (1997)
Phillips & Shine (2006)
Response to competitor
Various grasses
Sporobolus airoides
Alliara petiolata (Fig. 1)
Alkali sacaton grass
(invasive) garlic mustard
Reduced sensitivity to Centaurea allelopathy
Higher performance with invasive Acroptilon
Allelopathy negatively correlated w/
conspecific density
Reduced sensitivity to Alliara allelopathy
20–30
25–80
150
<30
<80
c. 75
Callaway et al. (2005)
Mealor & Hild (2007)
Lankau (2012)
<150
<150
Lankau (2012)
Pilea pumila
Clearweed
13–80 generations (Table 3). For instance, Alliaria petiolata
produces allelopathic sinigrin that negatively affect understory
plants in areas where it has been introduced – populations of
a native competitor (Pilea pumila) from high-sinigrin areas
had higher fitness than did ‘na€ıve’ populations in the presence
of A. petiolata, but lower fitness when A. petiolata was absent,
indicating both the evolution of sinigrin resistance, and its
associated cost (Lankau 2012).
As in the case of responses to abiotic factors, it is important
to confirm that phenotypic changes are due to genetic change
and that this genetic change is adaptive. For all the examples
in Table 3, the observed phenotypic changes were confirmed
to be due to heritable (genetic or epigenetic) rather than plastic responses to different environments. In six species, the
changes were confirmed to be adaptive, conferring higher
fecundity or lower mortality in the appropriate environment
(Kiesecker & Blaustein 1997; Carroll et al. 2001; Sheldon &
Jones 2001; Mealor & Hild 2007; Lankau 2012). Fitness was
not measured in the remaining six studies, but changes were
such as would be expected to increase fitness, such as phenology matching host fruit availability (Filchak et al. 2000) or
induced shell thickening previously shown to reduce vulnerability to predators (Freeman & Byers 2006). In future studies,
it would be helpful to survey the frequency of adaptive
responses to invasive species within native communities, and
to identify what traits of invaders or natives may promote or
constrain adaptive responses.
The role of phenotypic plasticity
Phenotypic plasticity allows for changes in phenotype within a
single generation, and as such is expected to play an
important role in both invasion and responses to environmental change (Matesanz et al. 2010; Palacio-Lopez & Gianoli
2011). Contrary to earlier hypotheses, recent reviews (Matesanz et al. 2010; Godoy et al. 2011; Palacio-Lopez & Gianoli
2011) have not supported the idea that invasive species are
more plastic or that increased plasticity is necessarily advantageous (Davidson et al. 2011). Plasticity can permit survival or
establishment under suboptimal conditions and therefore
allow subsequent adaptation, so it is likely to be quite important both for successful invasion and for species responding to
global change (Matesanz et al. 2010). However, by alleviating
the negative effects of environmental change on fitness, plasticity weakens selection in the short term (Merila 2012). Most
of the phenotypic responses to climate change documented to
date can be attributed to plasticity rather than genetic changes
(Merila 2012). A meta-analysis found that human disturbance
(introduction of species, harvesting, pollution) tends to lead to
larger phenotypic change in populations than do non-anthropogenic processes over similar timescales, and that this difference is largely attributable to plastic responses (Hendry et al.
2008).
Phenotypic plasticity itself is a trait that can evolve. While
only a limited number of studies have compared reaction
norms between invasive populations and their native sources,
several have found changes in plastic responses. Of the 10
studies of invasive plants reviewed by Matesanz et al. (2010),
half showed an increase in plasticity in introduced populations (though in at least one case this was due to founder
effects) and within-population genetic variation in plasticity
has generally been found in the few studies that have
looked for it (Matesanz et al. 2010). However, changes in
plasticity are not necessarily adaptive and the fitness consequences of plasticity have seldom been examined. Some invasive species appear to undergo canalisation, where initial
plasticity transitions into closer genotype–phenotype relationships. For instance, the same species of snail exhibited highly
plastic behavioural responses when exposed to an invasive
crab for < 60 years, less plasticity after 110 years and highly
canalised withdrawal behaviour in the crab’s native range
(Edgell et al. 2009). These examples from invasive species
© 2014 John Wiley & Sons Ltd/CNRS
642 Emily V. Moran and Jake M. Alexander
suggest that native species may exhibit similar increases or
decreases in plasticity under the altered selection pressures
imposed by climate change (Chown et al. 2010; Matesanz
et al. 2010). Meta-analyses of data from invasive species
could be further exploited to test hypotheses about when
plasticity is most advantageous, when it is likely to evolve
and under what circumstances it helps or hinders evolution
of mean trait values.
WHAT LIMITS EVOLUTIONARY RESPONSES TO
ENVIRONMENTAL CHANGE?
Time
One of the constraints on evolution most frequently considered in the context of climate change is time (Davis et al.
2005; Jump & Penuelas 2005). Invasive species demonstrate
that organisms facing altered environmental conditions can
evolve rapidly and in a variety of ways, but suggest that we
are unlikely to see substantial evolutionary change in ecologically relevant traits in fewer than 25 generations (Tables 2
and 3). In other words, all else being equal the species most
likely to exhibit substantial evolutionary responses to environmental change occurring over the next century may be those
with a generation time of 4 years or less. However, because
invasive species are often not studied until they are relatively
widespread, it is possible that substantial evolutionary change
occurred in < 25 generations in some species, but has simply
not been measured.
Genetic diversity
There has recently been much interest in the role of genetic
diversity and gene flow in determining species’ responses to
climate change (Kremer et al. 2012), and this is another area
in which invasive species can be informative. Founding populations of invasive species are often subject to genetic bottlenecks; as a result, neutral genetic diversity and heterozygosity
are lower in many invasive populations relative to their native
range (Dlugosch & Parker 2008a; Yue et al. 2010; Facon
et al. 2011). As the previous section shows, this does not
necessarily preclude either spread or local adaptation. Four
possible reasons include: the often weak correlation of neutral
and functional genetic diversity (Kohn et al. 2006); the purging of deleterious recessive alleles, which reduces inbreeding
depression (Facon et al. 2011); conversion of non-additive to
additive genetic variation (Prentis et al. 2008); and high population growth rates that limit the amount of variation lost
during a bottleneck (Huey et al. 2005).
However, this does not mean that genetic diversity is unimportant to adaptation and success. There is sometimes very
low variation in particular ecologically important traits even if
overall genetic diversity is high – for instance, tropical
Drosophila often exhibits low genetic variation and thus low
heritability in desiccation and heat tolerance (Merila 2012).
Some authors have suggested that a lack of genetic diversity
or a need to adapt to local conditions could be partly responsible for the long lags between establishment and spread in
some invasive species (Ellstrand & Schierenbeck 2000), and
© 2014 John Wiley & Sons Ltd/CNRS
Review and Synthesis
while direct evidence for this is scarce, indirect evidence of the
potential importance of genetic diversity and/or adaptation is
mounting (Crooks 2005). Dlugosch & Parker’s review (2008a)
suggests a temporal trend in genetic diversity, whereby
diversity is recovered due to multiple introductions and interpopulation gene flow. Admixture between different source
populations can significantly increase fitness in invasive
species (Keller & Taylor 2010; Zhang et al. 2010). Of the 13
species listed in Table 2 as examples of apparent adaptation
to climate, six appear to have been introduced from a single
source population (Weber & Schmid 1998; Huey et al. 2000;
Quinn et al. 2001; Lounibos et al. 2003; Dlugosch & Parker
2008b), while five likely derived from multiple source populations (Leger & Rice 2007; Etterson et al. 2008; Montague
et al. 2008; Keller et al. 2009). Rabbits were introduced to
Australia once, but the 13 founders were chosen from a
variety of wild and domestic sources (Zenger et al. 2003). The
original source of Echinochloa crus-galli in North America is
unknown, but there do not appear to have been any recent
introductions (Roy et al. 2000). These results from invasions
suggest that, for native taxa facing global change, efforts to
increase or maintain genetic diversity within native populations via translocations (Weeks et al. 2011) may indeed be
helpful, though not always necessary, in promoting adaptation. In the context of climate change, the direction of gene
flow is likely to be important, as we will discuss in a later
section.
Interspecific hybridisation appears to have aided invasive
spread in many plant taxa (Schierenbeck & Ellstrand 2009)
and even some animals (Nolte et al. 2005). In some cases
hybridisation may function much like admixture between populations (Ellstrand & Schierenbeck 2000; Zalapa et al. 2010),
but it can also lead to ‘transgressive’ phenotypes that allow
occupation of novel habitats (Ellstrand & Schierenbeck 2000;
Lambrinos 2004; Nolte et al. 2005; Rieseberg et al. 2007).
Climate change is already leading to increased hybridisation
in some taxa. For example, warmer temperatures have facilitated hybridisation between a southern Papilio butterfly with
multiple generations per year and a northern univoltine
species; the hybrid emerges late in the year and is evolving
different host plant use (Mercader et al. 2009). Other global
changes, particularly habitat alteration, can lead to increased
contact and hybridisation between native species. The consequences can be difficult to predict. In some cases, such introgression is arguably beneficial: Sculpin hybridisation enabled
occupation of novel, human-modified river habitat (Nolte
et al. 2005). However, hybridisation can also reduce fitness or
lead to the loss of distinct taxa (Genovart 2009; Muhlfeld
et al. 2009).
Conflicting selection pressures on the same trait
Multiple selection pressures acting on a trait in different directions can prevent changes in the trait from occurring or lead
to maladaptation with respect to a particular selection factor.
This could occur in populations exposed simultaneously to
changes in climate as well as other global change factors. One
global change factor that is likely to interact with climate
change in terms of species’ evolutionary responses is habitat
Review and Synthesis
fragmentation. Fragmentation tends to decrease the probability of successful dispersal, which reduces range shift velocity
(Jump & Penuelas 2005). If survival in the surrounding matrix
is very low, ‘stay-at-home’ phenotypes will be favoured even if
the initial colonists of a habitat patch were good dispersers,
and over time this selection pressure can reduce dispersal ability. Such reductions have occurred in weedy plants over < 15
generations, as documented in species growing in an urban
landscape or on oceanic islands (Lambrinos 2004; Cheptou
et al. 2008). Reduced gene flow makes the depletion of local
genetic diversity by selection and drift more likely, and this
could cause the rate of adaptation to increasingly lag behind
the rate of environmental change, increasing the risk of
extinction (Jump & Penuelas 2005). If the highly fragmented
nature of many landscapes today selects against dispersal, this
would tend to reduce the capacity for range shifts following
climate change, weaken demographic rescue effects and
increase genetic isolation. Further studies of invasive species
in fragmented vs. more continuous habitat could help to
reveal the role of such conflicting selection pressures in
shaping dispersal ability. For instance, it is currently unclear
what level of fragmentation (or matrix unsuitability) is
required for selection to favour reduced dispersal.
Correlations among traits under selection
Population genetics theory suggests that one important factor
limiting adaptation to novel environments may be correlations
between traits that are under selection (Guillaume 2011;
Merila 2012). For instance, invasive purple loosestrife has
developed a cline in flowering time matching changing growing season length along a latitudinal gradient, but there is a
negative correlation between early flowering and fecundity
that may constrain further northward spread by limiting seed
production in northern populations (Colautti et al. 2010).
Similarly, correlations between traits such that the major axes
of variation are not aligned with the selection gradients could
impede evolutionary responses of native species to climate
change. For example, evolution in a prairie plant in response to
climate change is likely to be constrained by trait correlations
(Etterson & Shaw 2001). However, the role trait correlations
play in niche expansion vs. stasis has not been fully examined in
either native or invasive species (Merila 2012). Studies of
evolution in invasive species often measure multiple phenotypic
traits (including size, fecundity and phenology), which could be
used to test hypotheses about the role of genetic trait
correlations, but this resource has been underused.
Population-level vs. species-level adaptation
While local adaptation may be important for populations
throughout a species’ range as climate warms (Jiguet et al.
2010), it is likely to be particularly important for populations
at the trailing edge (e.g. at low latitude). Here, populations
will be exposed to conditions that are not only locally novel,
but that are more extreme than those experienced by populations anywhere else in the species’ range (Fig. 2). If the
species’ distribution limit reflects limits to adaptation
(Soberon 2010), such populations will need to evolve new spe-
Invasion and evolution to global change 643
cies-level climatic niche limits in order to persist. Some have
argued that this is unlikely, as evidenced by the apparent
stability of species’ thermal limits in the fossil record (Parmesan 2006). Here again invasive species are instructive.
Whilst invasive species have demonstrated an impressive
ability to adapt to local climatic conditions over short time
periods (Table 2), evidence for the evolution of species-level
physiological tolerances is much more limited, and primarily
comes from the use of correlative niche models (also known
as climate envelope or species distribution models) that test
whether invasive populations occupy areas with combinations
of climatic variables that are not occupied in the native range
– in which case they are said to display a niche shift or expansion (Pearman et al. 2008). Results of such tests are mixed.
For instance, Gallagher et al. (2010) found that 77% of 26
invasive European plants in Australia showed evidence of a
niche shift or expansion, while Petitpierre et al. (2012), using
a different method, found that only 14% of 50 invasive plants
transferred between Europe and North America show more
than a 10% niche expansion.
However, apparent niche expansions are often hard to interpret. For example, a species may occupy environmental conditions that simply are not found in its native range, in which case
it is not clear whether the fundamental niche is broader than
previously suspected, whether the expansion is due to changes
in biotic interactions or if there has been evolution of physiological limits (Alexander & Edwards 2010). Methods that account
for whether the range of environmental conditions overlap
between ranges find more support for the conservation of native
range niche limits in the invasive range (Petitpierre et al. 2012).
Furthermore, the choice of environmental variables examined
(Peterson & Nakazawa 2008; R€
odder & L€
otters 2009) and the
use of presence–absence vs. physiological data (R€
odder et al.
2009) can affect whether a niche shift is observed.
The experimental tests required to disentangle these explanations are surprisingly scarce. One study found that nonnative populations of Lactuca serriola have expanded their
climatic niche substantially relative to their ancestral populations, and evolved clines in flowering phenology along a
climate gradient that matched similar clines in the native
range, but remain within the climatic niche of the species as a
whole (Alexander 2013). To our knowledge, only one experiment has yet tested whether invasive populations that inhabit
novel climates have evolved different physiological limits. Hill
et al. (2013) demonstrated that an apparent niche shift in the
mite Halotydeus destructor, identified through correlative
niche models, corresponded to an increased upper thermal
threshold and faster recovery from cold stress relative to
native populations. This shift in the fundamental niche
occurred within 40 years. If true niche expansions are rare (or
occur very slowly), even in invasive species, this would suggest
that rapid evolution in response to climate change is unlikely
to occur outside of the range of conditions experienced by a
species as a whole. On the other hand, if other invasive taxa
have expanded their physiological limits as H. destructor has
done, then evolution may be able to rescue ‘trailing edge’
populations of similar species threatened by climate change.
This information could also help inform our understanding of
likely responses to other global changes.
© 2014 John Wiley & Sons Ltd/CNRS
644 Emily V. Moran and Jake M. Alexander
(a)
Before rapid climate change
Review and Synthesis
(b) After rapid climate change
Low gene
ow
favours
local
adaptation
across
species
range
Gene ow
favours
local
adaptation
to new
climate
Climate envelope shifts to higher latitude
Selection for
increased
dispersal
ability
Selection on
species-level
climatic tolerance
Figure 2 Predicted shifts in selective pressures due to climate change. The size of the arrows indicates the favourableness of gene flow. Under current
climate (a), if populations are locally adapted, selection is expected to favour modest levels of gene flow among populations (arrows). As climate becomes
warmer (b), populations are no longer in equilibrium with their climate envelope. Selection now favours greater gene flow (bold arrows) among
populations that occupy a similar climate (which increases genetic variation), and from warmer-climate populations (bringing pre-adapted alleles), but gene
flow from colder-climate populations is maladaptive. Rear-edge populations must adapt to conditions that are beyond the species’ previous climatic
envelope, while leading-edge populations may experience selection for increased dispersal to reach newly suitable habitat (red arrow). All populations may
need to evolve to interact with new species that are encountered following range shifts.
TO WHAT EXTENT ARE INVASIONS A USEFUL MODEL
FOR EVOLUTIONARY RESPONSES TO
ENVIRONMENTAL CHANGE?
While there are many parallels between the selective pressures
faced by introduced species and those experiencing rapid environmental change, there are several reasons why invasive species might not reliably predict evolutionary responses in many
native species. Most notably, studies of evolution in invasive
species have focused on those that are successful, but the literature suggests that of all the species introduced into a given
region only a few successfully establish (or ‘naturalize’) and
an even smaller subset become invasive (Caplat et al. 2013).
Thus, data from invasive species may be biased towards species that show adaptive responses because those that do not
may be less likely to establish and spread. Studies of species
in earlier stages of naturalisation and spread would be helpful
in addressing this potential source of bias.
The novel biotic and abiotic selection pressures experienced
by invasive species are not completely analogous to the
© 2014 John Wiley & Sons Ltd/CNRS
changes in selection pressures predicted for native species
under environmental change. Native populations facing environmental change will rarely be suddenly exposed to the novel
environment – conditions will change gradually over time,
and so selection pressures are likely to be initially weaker.
This could increase the probability of evolutionary responses,
since lower population growth rates are required to avoid
extinction when the optimum trait value changes slowly
(Gomulkiewicz & Houle 2009). On the other hand, the relaxation of selection pressures from natural enemies may have
helped some invasive species to respond to strong abiotic
selection (Hornoy et al. 2011). Native species are not expected
to benefit from long-term enemy release (though transient
release could aid in range shifts; Moorcroft et al. 2006), so
responses to strong abiotic selection pressures might be more
constrained than during invasion.
Invasive species are often assumed to have higher reproductive output or population growth rates than many native species, which would tend to increase responses to selection.
However, evidence for this is mixed. One study comparing
Review and Synthesis
invasive and native plant populations found that invasive species had a higher population growth rate on average (Ramula
et al. 2008), while another found that the range of population
growth rates exhibited by native and invasive plants with similar life histories in the same area did not differ (Meiners
2007). Nevertheless, native species could exhibit slower adaptation to global change if environmental degradation has
already lowered their survival or reproductive success
(Gomulkiewicz & Houle 2009; Lavergne et al. 2010).
Severe population bottlenecks are less likely for species
experiencing climate change than for invasive species. However, species that have undergone relatively recent range
expansions often exhibit a gradient of progressively lower
genetic diversity from the original core distribution towards
the new range edge, resulting from sequential founder effects
(Huchon et al. 1999; Hampe & Petit 2005). In native populations, admixture between widely geographically separate
lineages (in the absence of direct human intervention) is not
likely to occur. Such admixture has historically been observed
to be advantageous in the invasive range but disadvantageous
in the native range (Verhoeven et al. 2010). This is most likely
because under stable environmental conditions the benefits of
maintaining locally adapted allelic combinations outweigh any
costs of inbreeding depression, whereas novel allelic combinations become more advantageous under novel selection pressures (Verhoeven et al. 2010). Several theoretical studies have
suggested that dispersal from large range-center populations
could push edge populations away from their phenotypic optimum, but other theoretical and empirical studies suggest that
the positive effect of increased genetic diversity tends to outweigh the negative effects of gene flow in ‘sink’ environments
(Lavergne et al. 2010). The changing selection pressures associated with climate change are likely to increasingly favour
admixture among populations from climatically similar areas
and gene flow from warm-climate populations to (historically)
cool-climate populations (Fig. 2; Kremer et al. 2012). Gene
flow from cool-climate to warm-climate populations, however,
is likely to produce maladapted phenotypes and will probably
not be favourable unless the benefits of increased diversity
outweigh the costs of more cold-adapted alleles (Kremer et al.
2012). This latter situation is likely similar to the dispersal
from ‘core’ areas experienced by populations of invasive species establishing in areas outside the historical climate niche,
so these cases of putative niche expansion (or failure to
expand the niche) could also prove useful in examining the
role of dispersal in local adaptation.
These examples suggest that rapid evolution might be more
likely to occur during invasions than for many native species
responding to climate change. One area in which the reverse
could be true is adaptation to novel hosts or mutualists that
are encountered following range shifts. Host shifts are more
likely to occur with decreasing taxonomic distance between
hosts (Carroll 2007). Species often have multiple close relatives within a continent, so it is likely that at least some insect
species will be able to shift their ranges even if their original
hosts do not, by taking advantage of similar hosts in areas
that are becoming climatically suitable (Thomas et al. 2001).
Conversely, this suggests that range-shifting species may not
benefit for long from enemy escape.
Invasion and evolution to global change 645
FUTURE DIRECTIONS: HOW INVASIVE SPECIES CAN
FURTHER INFORM OUR UNDERSTANDING OF
EVOLUTIONARY RESPONSES TO ENVIRONMENTAL
CHANGE
The factors affecting the likelihood and extent of an evolutionary response include a) the strength of selection, b) the
amount local genetic variation and gene flow, c) demographic
traits such as lifespan, generation time, and fecundity, and d)
genetic correlations in the traits under selection. While demographic traits are relatively well known for many species, estimating the other quantities usually requires manipulative
experiments. As a result, it is unlikely that we will be able to
obtain estimates for all species of concern. However, by examining a broad range of species including both native species
exposed to current environmental change and introduced species that have adapted to new environments, it may be possible to draw generalisations about what kinds of populations
or species exhibit ecologically relevant evolutionary change
under particular types of environmental change, what types of
change are more likely and what constraints tend to limit evolutionary responses.
We have highlighted ways in which invasive species can
inform us about evolutionary responses of native species to
global change. However, this analogy could be more fully
exploited when designing studies of evolution during invasion,
to enable additional insights into the role of evolution in species’ responses to environmental change in general. We believe
that greater collaboration between invasion biologists and global change researchers would be beneficial for both fields. In
this section we suggest several outstanding research questions
for which invasive species would provide useful information.
How will selection pressures due to changing biotic and abiotic
factors interact?
Invasive species sometimes experience relaxed selection from
natural enemies when they leave specialist enemies behind in
the native range, which might increase their ability to respond
to other selection pressures (Callaway & Maron 2006; Hornoy
et al. 2011). It is important to assess the extent to which abiotic and biotic selection pressures interact and constrain evolutionary responses in order to predict how native species
might respond to diverse drivers of global change, especially
because native species may be less likely to benefit from
enemy release than non-natives. Studies of adaptation in the
presence vs. absence of enemies (e.g. biocontrol agents) could
help disentangle these factors.
Do species attributes predict the likelihood of evolutionary
responses to environmental change?
If invasive species with particular combinations of life-history
or demographic traits exhibit stronger evolutionary responses
to novel environments, this could enable us to predict which
native species may show the strongest evolutionary responses to
global change. Furthermore, while invasive species are often
assumed to have higher fecundity or population growth rates
than native populations of the same or related species, this has
© 2014 John Wiley & Sons Ltd/CNRS
646 Emily V. Moran and Jake M. Alexander
seldom been tested (but see Ramula et al. 2008; Meiners 2007);
such comparative and meta-analytical studies would help to
determine whether or not data from invasive species are likely
to overpredict the potential for evolutionary responses.
How much relevant functional genetic diversity exists within
populations and species?
The ability of populations to adapt to rapid environmental
change will be limited by the availability of functional genetic
variation on which selection can act. However, we still know
very little about the amounts of additive genetic variation in
ecologically important traits (Dlugosch & Parker 2008a), and
how this variation is distributed across a species’ range in
either native or invasive species. Comparisons of functional
diversity in introduced populations (that do or do not evolve)
to that of native populations could illuminate what levels of
diversity are required and the effect of bottlenecks, admixture
and gene flow on this diversity.
How will climate change and habitat fragmentation interact to
affect dispersal ability?
The availability of new suitable habitat and habitat fragmentation are likely to have conflicting effects on dispersal ability.
Given the highly fragmented nature of many modern landscapes, it is important to understand how these two factors
will interact to shape species’ ability to shift their ranges in
response to climate change and to exchange genes between
populations. Invasive and weedy species could be a useful
model for determining what combinations of patch distance,
initial dispersal ability and matrix quality select for or against
high dispersal ability.
Can evolution rescue populations at the trailing edge of a species’
distribution?
Although many of the apparent niche shifts in invasive species
are towards warmer or drier conditions (Broennimann et al.
2007; Loo et al. 2007; Compton et al. 2010; Petersen 2012), it is
not known whether this is a result of evolutionary adaptation
(but see Hill et al. 2013). Comparative studies of native and
non-native populations collected from the warm edge of their
climate niche in each range could be used to test the lability of
species-level climatic tolerances, by rearing accessions across a
climatic gradient or testing physiological limits in the lab (e.g.
Alexander et al. 2012; Hill et al. 2013). This would be especially
valuable for non-native species that show evidence of niche
shifts. The frequency with which invasive species are able to
expand their fundamental niche will give insight into the ability
of evolution to rescue ‘trailing edge’ native populations.
SUMMARY
Data from invasive species complement data on evolutionary
responses in native species. Studies of invasive species can
examine evolutionary change over a wide range of timescales
in a variety of taxa, common garden data on multiple traits
are often available, and there are opportunities to study the
© 2014 John Wiley & Sons Ltd/CNRS
Review and Synthesis
relative importance of biotic and abiotic limiting factors.
However, data on the earliest phases of establishment and
adaptation are generally lacking, and the genotype/phenotype
of founders must usually be inferred through comparisons
with current native-range populations. Conversely, studies of
native species often involve longitudinal studies that yield
information about both the selective pressure and the
response (Merila 2012), but in the case of recent environmental changes affecting species with long generation times it may
be too soon to detect genetic changes.
Studies of invasive species suggest that rapid evolution in
response to the novel abiotic and biotic conditions produced
by global change could occur rapidly (25–150 generations),
but that the potential for adaptation to conditions outside the
current species-level climatic niche may be more limited. Holt
(1990) noted that species that are theoretically most likely to
exhibit strong adaptive responses that could promote their
persistence under climate change are also those that are
already at lower risk of extinction – for example, those with
high abundances, short generation times and/or high dispersal
rates. The empirical data from invasive species that have accumulated over the past 20 years support these hypotheses, as
well as the predictions from population genetics theory that
evolutionary responses to novel environments will be constrained by generation time, lack of additive genetic variation
or correlations between traits under selection. However, studies on invasive species suggest that the level of neutral genetic
variation, commonly used as a proxy for overall genetic variation, may not always be a good predictor of a species’ ability
to adapt to changing conditions.
Data from invasive species may tend to overestimate the
rate of evolutionary responses to climate change if, in the
non-native range, (1) climatic selection pressures are stronger,
(2) biotic selection pressures are relaxed, (3) population
growth rates are higher, or (4) genetic admixture between populations is more common. Conversely, native species may
show faster responses if local genetic diversity is higher or if
the presence of related species of mutualists or hosts within a
continent facilitates the formation of new biotic interactions.
However, knowledge about the relative strength of selection
from local environmental change vs. translocation to a novel
environment (as in the case of invasive species), the amount of
functional genetic variation in climate-related traits or population growth rates in native vs. non-native ranges is still lacking, particularly for animals. Further research addressing these
factors, as well as the frequency of true evolutionary niche
expansion in invasive species, would be helpful in understanding the potential for evolutionary responses to global change.
ACKNOWLEDGEMENTS
We thank Jonathan Levine, Jeff Diez and Nicky Lustenhower, the editors of this journal, and all reviewers for their comments on this manuscript.
AUTHORSHIP
EVM wrote the initial draft and collected most of the references. JMA substantially restructured the MS, contributed
Review and Synthesis
additional references and created the figures. Both authors
contributed substantially to subsequent revisions.
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Editor, Frederick Adler
Manuscript received 14 October 2013
First decision made 23 November 2013
Second decision made 24 January 2014
Manuscript accepted 30 January 2014
© 2014 John Wiley & Sons Ltd/CNRS