Download - Wiley Online Library

Journal of
Ecology 2006
94, 1233–1248
ESSAY REVIEW
Blackwell Publishing Ltd
The rough edges of the conservation genetics paradigm
for plants
N. J. OUBORG, P. VERGEER* and C. MIX
Department of Molecular Ecology, Radboud University Nijmegen, Toernooiveld 1, 6525ED Nijmegen, The Netherlands
Summary
1 Small and isolated populations of species are susceptible to loss of genetic diversity,
owing to random genetic drift and inbreeding. This loss of diversity may reduce the
evolutionary potential to adapt to changing environments, and may cause immediate
loss of fitness (cf. inbreeding depression). Together with other population size-dependent
stochastic processes, this may lead to increased probabilities of population extinction.
2 This set of processes and theories forms the core of conservation genetics and has
developed into the conservation genetics paradigm. Many empirical studies have concentrated on the relationship between population size and genetic diversity, and in many
cases evidence was found that small populations of plants do indeed have lower levels of
genetic diversity and increased homozygosity. Although less empirical attention has
been given to the relationship between low genetic diversity, fitness and, in particular,
evolutionary potential, the paradigm is now widely accepted.
3 Here we present five areas of the paradigm which could be refined, i.e. the ‘rough’
edges of the conservation genetics paradigm.
4 Treating population size and isolation not as interchangeable parameters but as
separate parameters affecting population genetics in different ways could allow more
accurate predictions of the effects of landscape fragmentation on the genetic diversity
and viability of populations.
5 There is evidence that inbreeding depression may be a genotype-specific phenomenon,
rather than a population parameter. This sheds new light on the link between population
inbreeding depression and the expected increased probability of extinction.
6 Modern eco-genomics offers the opportunity to study the population genetics of
functional genes, to the extent that the role of selection can be distinguished from the
effects of drift, and allowing improved insights into the effects of loss of genetic diversity
on evolutionary potential.
7 Incorporating multispecies considerations may result in the generally accepted
notion that small populations are at peril being called into question. For instance, small
populations may be less capable of sustaining parasites or herbivores.
8 Comparative studies of endangered, common and invasive species may be a valuable
approach to developing conservation biology from a phenomenological case study
discipline into one investigating the general principles of what sustains biodiversity.
9 The issues discussed set an agenda for further research within conservation genetics and
may lead to a further refinement of our understanding and prediction of the genetic effects
of habitat fragmentation. They also underline the need to integrate ecological and genetic
approaches to the conservation of biodiversity, rather than regarding them as opposites.
Key-words: dispersal, eco-genomics, habitat fragmentation, invasions, multispecies
interaction
Journal of Ecology (2006) 94, 1233 –1248
doi: 10.1111/j.1365-2745.2006.01167.x
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society
Correspondence: N. J. Ouborg (tel. +31 24 3652470; e-mail [email protected]).
*Present address: School of Biology, University of Leeds, Leeds LS2 9JT, UK
1234
N. J. Ouborg,
P. Vergeer &
C. Mix
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
Introduction
Ecology is the scientific study of the interactions that
determine the distribution and abundance of organisms
(Krebs 1972), and ecologists try to understand the
processes that influence biodiversity. In the modern
landscape, biodiversity is endangered by the deteriorating quality of the environment and by a continuous
decrease in the amount of space that is available for the
natural environment. These two elements of biodiversity
threat form the basis of two models of nature, which
could be called paradigms in Kuhn’s (1996) terminology.
These paradigms underlie the interpretation and
appreciation of biodiversity problems and their
possible solutions.
The first paradigm, the ‘habitat quality’ paradigm,
emphasizes that biodiversity problems are the consequence of habitat quality changes, leading to loss of
species or populations that are not able to adapt to
these changes. The solution to biodiversity problems
within this paradigm lies in the control of habitat quality:
management strategies to preserve the high quality of
habitats where it is still present or measures to restore
this quality where the habitat has deteriorated.
The second paradigm is the ‘conservation biology’
or ‘conservation genetics’ paradigm (as we concentrate
mainly on the genetic part of conservation biology, we
use the latter term throughout the rest of this paper). In
this paradigm, biodiversity problems are appreciated
in the light of characteristics of populations (or species),
rather than in that of abiotic habitat characteristics:
populations of species are endangered as a result of their
small size and high degree of isolation from conspecific
populations. Within this paradigm, solutions to biodiversity problems are defined in terms of population
reinforcement (for instance by artificially increasing
the number of individuals in small populations) or
reduction of isolation (for instance by establishing
corridors between existing conspecific populations).
Although these two paradigms do present two alternative interpretations of nature, they are not incompatible
or mutually exclusive. The habitat quality paradigm
does not deny that small populations may be more prone
to extinction, but regards these problems as secondary
to the problems imposed by habitat deterioration. By
contrast, the conservation genetics paradigm fully
acknowledges the importance of habitat quality, but
emphasizes that even in optimal habitats, populations
may be threatened with extinction. Shifting from the
habitat quality paradigm to the conservation genetics
paradigm implies shifting from a habitat-orientated to
a genetics-orientated approach, from an ecosystemorientated to a population-orientated view of nature
and from a deterministic to a stochastic treatment of
biological processes. Moreover, the distinction between
these paradigms meets at least one criterion for a paradigm shift, as defined by Kuhn (1996): each paradigm is
represented by its own community of scientists, united
in their own societies (the Society for Restoration
Ecology and Society for Conservation Biology, respectively), which tend to publish in their own journals.
In the rest of this paper, we concentrate on the conservation genetics paradigm. First, we outline the main
model of nature in this paradigm, and then discuss five
areas of the paradigm (‘the rough edges’) that could be
refined. These discussions contribute to our strong
conviction that optimal understanding of how biodiversity is threatened, and how it can be conserved, is
most efficiently achieved when the two paradigms are
integrated, rather than being regarded as opposites.
The conservation genetics paradigm
In their pivotal 1981 book Conservation and Evolution,
marking the birth of the discipline of conservation
biology, Frankel and Soulé presented a scientific
framework showing how evolution and the dynamics
of genetic variation within and among populations are
important for the conservation of endangered species.
Their point of departure was the fact that many endangered species only occur in small, isolated populations,
resulting in the increased impact of a series of stochastic
processes that negatively influence the chance of survival
of these populations. More specifically, when a population becomes small, its growth rate is increasingly
affected by four types of stochastic processes (Fig. 1).
First, demographic stochasticity – the random fluctuation in demographic parameters among individuals
within a population – will lead to stochastic variation
of the population growth rate in small or very small
populations (Engen et al. 1998; Fox & Kendall 2002;
Drake 2005; Fig. 1-1). Second, stochastic variation in
environmental parameters over time and space will lead
to stochastic variation of key demographic parameters,
a phenomenon known as environmental stochasticity
(Fig. 1-2). This type of variation will have significant
impact on the population growth rate, even in averagesize populations (Lande 1993; Boyce et al. 2006). Third,
unpredictable catastrophes with a major impact on
population size, such as hurricanes, flooding and forest
fires, are more likely to lead to the extinction of small
than of large populations (cf. catastrophic stochasticity;
Goodman 1987; Lande 1993; Fig. 1-3). The fourth type,
genetic stochasticity, is the impact of random populationgenetic processes (i.e. genetic drift) on demographic
parameters and population growth rate, an impact that
increases with decreasing population size and has
attracted much attention over recent decades (Soulé &
Wilcox 1980; Schonewald-Cox et al. 1983; Soulé 1986;
Ellstrand & Elam 1993; Young & Clarke 2000; Frankham
et al. 2002).
The way in which genetic stochasticity, or genetic
erosion as it is frequently referred to (e.g. Cole 2003;
Oostermeijer et al. 2003), may affect extinction probabilities of populations is as follows (schematically
outlined in Fig. 1). Small populations will be highly
affected by genetic drift, as the extent of drift is an
inverse function of effective population size. Genetic
1235
The rough edges of
the conservation
genetics paradigm
Fig. 1 This figure schematically represents the two views on threats to biodiversity: the ‘habitat quality paradigm’ emphasizes the
left part of the scheme, whereas the ‘conservation genetics paradigm’ emphasizes the right part. Negative correlations are
represented by dotted lines and positive correlations by solid lines. Numbers in circles refer to the types of stochasticity: 1,
demographic stochasticity; 2, environmental stochasticity; 3, catastrophic stochasticity; genetic stochasticity is represented by the
right side of the scheme. Upper-case letters refer to the ‘rough edges’ discussed herein. A, population size vs. isolation; B,
inbreeding depression; C, neutral markers vs. functional genes; D, multispecies interactions. E, from phenomenology to general
principles, is represented by a comparison of diagrams for rare species with those for common and/or invasive species.
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
drift will ultimately lead to a loss of alleles from the
population, and will also bring about an increase in
homozygosity, averaged over loci and individuals
(Hedrick 2005a). Moreover, in small populations, individuals will become increasingly related to each other
over time (Frankham et al. 2002), leading to increased
bi-parental inbreeding and increased homozygosity.
Thus, small populations are expected to lose genetic
variation over time, which is expressed as a decrease in
the average number of alleles per locus and an increase
in the average homozygosity.
This may have three interrelated consequences.
First, increased homozygosity may be associated with
reduced fitness. Negative correlations between fitnessrelated characters and homozygosity have frequently
been reported (e.g. Mitton & Grant 1984; Allendorf &
Leary 1986; Mitton 1997). Second, small populations
are frequently associated with inbreeding depression.
Both dominance and overdominance models of inbreeding depression predict that increased homozygosity may
lead to reduced fitness (Charlesworth & Charlesworth
1999; Carr & Dudash 2003). Third, the loss of genetic
variation from a population may reduce its potential to
adapt to changing environments (Nunney & Campbell
1993; Frankham 1995). Rapid environmental changes
may therefore raise the probability of extinction in
small populations that lack the genetic variation
needed for an adequate evolutionary response (Bijlsma
& Loeschcke 2005).
These fitness aspects may increase the probability
that a population becomes extinct. This may be either
an immediate effect, when increased homozygosity and
inbreeding depression affect those life stages with a
high relative contribution to the population growth
rate (i.e. high elasticity; De Kroon et al. 1986), or a
postponed effect, when evolutionary potential is lost.
In addition to this increased extinction probability, the
reduced growth rate may result in a further decrease in
population size, making it a self-reinforcing process.
Once a population enters this process, its reinforcing
nature results in a vortex, ultimately leading to extinction
(e.g. Dennis 2002; Fagan & Holmes 2006; Hedrick
et al. 2006).
Isolation – the lack of interpopulation dispersal –
may play a role at both the demographic and the
genetic levels. Migration may lead to the demographic
rescue of small populations (Brown & Kodric-Brown
1977), with influx of individuals increasing the population growth rate. Gene flow, the genetic counterpart of
dispersal, may lead to the ‘genetic rescue’ of genetically
eroded populations, by reducing inbreeding depression
and increasing the level of genetic variation (Richards
2000; Ingvarsson 2001).
Together, this set of theories, hypotheses and relationships forms the basis of the conservation genetics
paradigm. Since the early 1980s, significant advances
have been made in conservation genetics theory. Topics
that have received considerable theoretical attention
include the role of mutation accumulation in small
populations (e.g. Lynch et al. 1995; Schoen et al. 1998;
Higgins & Lynch 2001), the role of inbreeding depression
in population extinction (e.g. Mills & Smouse 1994;
1236
N. J. Ouborg,
P. Vergeer &
C. Mix
Brook et al. 2002; Glemin et al. 2003; Armbruster &
Reed 2005), the possibilities and limitations of the
purging of genetic loads (e.g. Whitlock 2000; Crnokrak
& Barrett 2002; Glemin 2003) and the role of spatially
subdivided populations (i.e. metapopulations) (e.g.
Higgins & Lynch 2001; Hanski & Gagiotti 2004;
Bouchy et al. 2006; Theodorou & Couvet 2006) in
understanding and predicting the impact of genetics on
population or metapopulation extinction. Almost
without exception, these studies have underlined the
importance of genetics for conservation and have
provided a further basis for a model of nature that
emphasizes that small, isolated populations may be
threatened for genetic (as well as other) reasons.
The rough edges
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
In recent decades, the conservation genetics paradigm
has been embraced by population biologists, population geneticists and conservationists, and has become
the framework for empirical studies of rare and
endangered species (Barrett & Kohn 1991; Young et al.
1996; Young & Clarke 2000), the effects of landscape
fragmentation on population viability (Young &
Clarke 2000), the design of measures against habitat
fragmentation effects (Bullock & Hodder 1997; Van
Groenendael et al. 1998; Vergeer 2005), the conservation
of populations of plants and animals (Schonewald-Cox
et al. 1983; Ellstrand & Elam 1993) and the design of
breeding programmes in zoos (Frankham et al. 2002).
Many ecological studies on plants are now based on
some form of conservation genetics framework (e.g.
Dupré & Ehrlén 2002; Luijten et al. 2002; Booth &
Grime 2003; Lienert & Fischer 2003; Oostermeijer
et al. 2003; Vergeer et al. 2003; Ashworth et al. 2004;
Brys et al. 2004; Kolb 2005; Nicolé et al. 2005; Prentice
et al. 2006).
In spite of this wide acceptance, there is an ongoing
debate about the relevance of genetic stochasticity,
compared with the other types of stochasticity, for
extinction. It has been contended that populations go
extinct owing to demographic and environmental
stochasticity before genetic stochasticity becomes important (Lande 1988; Schemske et al. 1994; Holsinger
et al. 1999; but see Frankham & Ralls 1998; Spielman
et al. 2004). This debate partly reflects the ‘controversy’
between the habitat quality and conservation genetics
paradigms. Several reviews have compiled and evaluated
the empirical evidence for the importance of genetic
effects (e.g. Young et al. 1996; Booy et al. 2000), including
the very recent review by Leimu et al. (2006). In this
paper we follow a different approach. Instead of
evaluating the support for the conservation genetics
paradigm, we discuss five areas where we feel that
refinement of the paradigm is possible and needed
(Fig. 1A–E). As such, we do not aim to criticize the
core of the conservation genetics paradigm, but rather
try to refine parts of it, the parts we call the rough edges
of the paradigm. We do not claim that our discussion
will resolve the ecology-vs.-genetics debate; instead,
we argue that this debate is in fact asking the wrong
question: what matters is not the question of what is
more important for conserving biodiversity – ecology
or genetics – but how ecology and genetics interact in
affecting the viability of populations and species.
Population size vs. isolation
Habitat fragmentation and its effects on the persistence
of populations and species is of major concern to
conservation biology. When landscapes become fragmented, this may result in a decrease in the number of
populations, a decrease in the average size of remaining
populations and an increase in the average interpopulation distance. The conservation genetics paradigm includes these general effects by emphasizing the
impact of reduced population size on the persistence of
local populations, and by incorporating dispersal (or its
genetic counterpart gene flow) as a means to mitigate
these effects of reduced population size. In studying the
effects of habitat fragmentation, many empirical studies
have focused on the impact of reduced population size.
Although in most cases it was acknowledged that
isolation may play a role, isolation or dispersal were not
explicitly quantified or investigated. This may be due to
the inherent difficulties of defining isolation (Moilanen
& Nieminen 2002; Soons et al. 2005) or measuring
long-distance dispersal (e.g. Ouborg et al. 1999; Cain
et al. 2000). Nevertheless, there are reasons to be more
precise and use population size and isolation as two
separate parameters, each with their own expected
effects (Fig. 1A).
Fahrig (2003) pointed out that even though at the
landscape level fragmentation is expected to lead to
reduced population size and increased interpopulation
distances, this is just one of several possible scenarios.
She argued that, in fact, the number of remaining
patches, the average size of remaining patches and the
mean patch isolation may each either increase or
decrease. Moreover, these three parameters will respond
independently to landscape fragmentation and should
therefore not be treated as interchangeable when
studying the impact of habitat fragmentation.
Reduced population size and increased isolation will
affect the persistence of local populations in different
ways. From a genetic point of view, reduced population
size results in increased genetic drift within local populations, leading to random loss of alleles from these
populations. At the same time, the average degree of
relatedness among individuals in a small population
increases faster over time than in a large population,
leading to increased levels of bi-parental inbreeding,
increased homozygosity and, often, a reduction in
average fitness (Fig. 1).
By contrast, increased isolation, i.e. increased
average interpopulation distance, results in a decrease
in among-population dispersal and gene flow. This
may have several consequences. The demography and
1237
The rough edges of
the conservation
genetics paradigm
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
evolutionary dynamics of populations may become
more independent of each other. The balance between
genetic drift and gene flow may shift towards the former
and the genetic differentiation between populations
may increase (Wright 1969). The likelihood of genetic
rescue (Ingvarsson 2001; Hedrick 2005b) of populations
(i.e. the influx of ‘new’ genetic diversity into genetically
impoverished populations) may decrease, as may the
likelihood of demographic rescue (Brown & KodricBrown 1977). In a heterogeneous landscape, where
isolated patches differ in environmental parameters, local
selection may result in local adaptation of plant populations (Hufford & Mazer 2003; see below). Thus, not
only do the landscape characteristics lead to reduced
exchange of individuals (cf. seeds, pollen) but selection
enhances differentiation in genetic quality between
populations (as a consequence of local adaptation),
leading to a further increase in effective isolation. If the
situation involves a metapopulation, its genetic and
demographic structure changes (Ouborg & Eriksson
2004). Based on a meta-analysis of 25 studies on the
effects of habitat loss on species diversity in animals,
Bender et al. (1998) concluded that there is good evidence that migration plays an independent role in the
persistence of species. Whether a similar rescue effect
could be prominent in plants, which generally have
much lower migration rates and smaller migration
distances than animals, is less clear, but several studies
have demonstrated that isolation may affect extinction
regardless of local population size (e.g. Dupre & Ehrlén
2002; Hooftman & Diemer 2002; Krauss et al. 2004;
Matthies et al. 2004; Ozinga et al. 2005).
The current conservation genetics paradigm is
strongly population-centred and treats dispersal and
gene flow as forces from outside local populations, not
affected by the genetic erosion process within the
populations. Here we present three reasons to refine
this approach towards a more realistic incorporation of
dispersal and isolation.
First, dispersal is not just affected by landscape-level
processes, but may be affected by local genetic erosion
processes as well. Long-distance seed dispersal results
from the integration of a number of demographic
processes, starting with seed production per individual
and followed by movement of seeds through space,
germination and establishment. Each of these components might be affected by genetic erosion. Inbreeding
depression in fecundity, germination and establishment
has often been demonstrated in plants (e.g. Husband &
Schemske 1996; for an overview see Keller & Waller
2002). However, the movement of seeds might also be
affected by local population processes. Inbreeding
effects on seed size, seed weight and the structure of
dispersal-aiding appendages (plumes, bristles, etc.) are
complex (Donohue 1999). This is the consequence of
the complex genetic make-up of seeds: seed coats are
entirely maternal, the embryo is entirely a diploid
assembly of maternal and paternal genes, while the
endosperm is frequently a triploid mix of maternal and
paternal contributions. This makes it rather difficult to
investigate and demonstrate inbreeding effects on seed
size. Nevertheless, a few examples of such studies do
exist (Schmitt et al. 1985; Pico et al. 2004a; Mix et al.
2006)
In a study of the effects of habitat fragmentation in a
Dutch grassland landscape on the metapopulation
structure and gene flow of two plant species, Mix et al.
(2006) found that for the wind-dispersed Hypochaeris
radicata, terminal velocity of seeds – a characteristic
determining the dispersal potential of individual seeds
(Soons & Heil 2002; Tackenberg 2003; Tackenberg
et al. 2003; Soons et al. 2004) – was correlated with
population size. The smaller populations had lower
terminal velocities, indicating a higher dispersal potential (Fig. 2a). Although it was unclear whether these
field effects could be fully accounted for by direct inbreeding effects on seed size (Mix et al. 2006), a separate
Fig. 2 (a) Relationship between the logarithm of population
size and terminal velocity of seeds in field-collected seeds
from eight Dutch populations of Hypochaeris radicata. Each
population is represented by the mean (and standard error) of
measurements of eight seeds per individual, for 15 individuals
per population (after Mix et al. 2006). (b) Effect of inbreeding on the terminal velocity of seeds of H. radicata. Three
inbreeding levels were created in the glasshouse: a self-group
(seeds obtained from pollinating plants with their own
pollen), outcross1 (seeds from a cross between two different
individuals from the same population) and outcross2 (seeds
from a cross between two individuals from different populations). Bars (and standard error) indicate the average effect
on 60 families, each family and pollination treatment being
represented by eight seeds (after Mix et al. 2006).
1238
N. J. Ouborg,
P. Vergeer &
C. Mix
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
glasshouse experiment with H. radicata demonstrated a
significant effect of inbreeding on terminal velocity
(Fig. 2b). Higher inbreeding levels resulted in higher
terminal velocities, and thus in a reduced dispersal
potential. Although field effects and glasshouse results
thus pointed in different directions, these results at
least indicate that the dispersal of seeds, a regional
process, may be affected by local within-population
processes. More experiments are needed in other landscapes and with other species to investigate the general
applicability of these results. At the very least, however,
these results suggest that isolation may be at least
partly affected by local genetic erosion, rather than by
a regional process independent of local processes.
Second, the conservation genetics paradigm places
strong emphasis on the effects of genetic drift and inbreeding. However, demographically important traits
will also be influenced by natural selection. Although it
is generally assumed that in small populations drift
might override the effects of selection (note, however,
that experimental evidence for this is very scant), a trait
such as dispersal potential is known to be selected for
or against in a regional population or metapopulation
context (McPeek & Holt 1992; Olivieri et al. 1995; Leimar
& Norberg 1997; Olivieri & Gouyon 1997; Travis &
Dytham 1999; Ronce et al. 2000, 2005). As the average
interpopulation distance increases, dispersing seeds
will more often be lost to the inhospitable surrounding
landscape. This implies that at regional scales, there
will be increased selection against dispersal. Evidence
for this was presented in a study by Cody & Overton
(1996), who found reduced dispersal potential in relatively isolated plant populations on Vancouver Island.
Colas et al. (1997) demonstrated that fragmented
populations of Centaurea corymbosa had very low dispersal capacity, and argued that even though unoccupied
but suitable sites were present nearby, the low dispersal,
and the resulting selection against dispersal capacity,
would eventually drive the species to extinction in this
area. Mix et al. (2006) found a correlation between the
terminal velocity of seeds of Succisa pratensis and the
distance to the nearest neighbouring population
(Fig. 3), indicating that less isolated populations had
higher dispersal potential. The final outcome of selection
for dispersal will be a balance between costs (i.e. loss of
individuals due to dispersal) and benefits (i.e. genetic
rescue) at both local and regional scales. This balance
between costs and benefits of dispersal deserves more
empirical attention within the conservation genetics
paradigm.
Third, dispersal also plays a crucial role in determining
the potential for differentiation between populations.
In a heterogeneous environment, selection may result
in local adaptation of genotypes (Linhart & Grant
1996). The extent of local adaptation, a process of great
importance for plants with their sedentary habit,
depends on the level of exchange of genes between the
various habitats. The current conservation genetics
paradigm tends to ignore the impact of selection and
Fig. 3 Relationship between population isolation (measured
as the logarithm of the distance to the nearest neighbouring
population) and terminal velocity of field-collected seeds of
eight Dutch populations of Succisa pratensis. Each
population is represented by the mean (and standard error) of
eight seeds per individual, for 15 individuals per population
(after C. Mix et al. unpublished data).
local adaptation. Dispersal is regarded as a beneficial
process, both because the paradigm focuses on the persistence of local populations and because it emphasizes
the effects of drift and inbreeding. From this perspective,
dispersal leads to an influx of genetic material, reversing
the loss of alleles, promoting heterozygosity and possibly leading to heterosis (i.e. the increased fitness of
offspring of crosses between parents from different
populations). But if different populations are differently
adapted in a heterogeneous landscape, dispersal might
disrupt this local adaptation, resulting in a decrease
in local average fitness (i.e. outbreeding depression)
rather than an increase. Examples of both heterosis
and outbreeding depression have been found (see the
overviews by Hufford & Mazer 2003 and Tallmon et al.
2004). The topic of local adaptation is starting to draw
the attention of applied conservation biologists, who
are considering management practices such as introduction or reintroduction and ecological restoration
(Hufford & Mazer 2003; Tallmon et al. 2004; Vergeer
et al. 2004). The topic is also highly relevant to the
discussion of what to preserve, as preserving several
smaller, isolated populations might maintain more
ecologically and evolutionarily important genetic
variation at a regional scale than preserving one large
population (see, for example, the SLOSS debate:
McCarthy et al. 2005).
In summary, the conservation genetics paradigm
should be refined by incorporating dispersal in a
more realistic way. Investigating the specific impact of
dispersal on local and regional population and metapopulation persistence, by adopting an evolutionary
perspective on dispersal and by incorporating the
effects of selection in heterogeneous environments,
might significantly improve the adequacy of the paradigm. Studies are needed to evaluate the adequacy of
various measures of population isolation, and to study
1239
The rough edges of
the conservation
genetics paradigm
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
the balance between genetic rescue and outbreeding
depression in plants.
Inbreeding depression: population parameter or
genotypic trait?
Inbreeding depression is defined as the reduction
in fitness in inbred individuals as compared with
their outbred relatives (Charlesworth & Charlesworth
1987). Genetic drift and bi-parental inbreeding will be
higher within small populations than within large
populations. Both processes result in an increased
average homozygosity, and increased homozygosity
allows the expression of recessive deleterious alleles,
resulting in inbreeding depression (Charlesworth &
Charlesworth 1999). Therefore, inbreeding depression
plays an important role in the conservation genetics
paradigm (Fig. 1B). Within the current paradigm,
the degree of inbreeding depression is treated as a
population parameter. In general, research to establish
this parameter involves individuals from a particular
population being both selfed and outcrossed. The
offspring of both crossing types is then raised and the
difference in fitness between the two offspring types is
summarized in a parameter, δ, the population inbreeding
depression. Alternatively, crosses can be made to create
a number of inbreeding levels (e.g. selfings, sib-mating
and unrelated crosses), after which inbreeding level is
regressed on the fitness of the resulting offspring and
the population inbreeding depression is expressed as
the (standardized) regression coefficient. These methods
of measuring inbreeding depression yield the potential
decrease in fitness if the inbreeding level in the population
increases over time as a consequence of continued small
population size. The current inbreeding depression, or
reduced fitness, is often measured by regressing population size on average fitness per population.
This method of incorporating inbreeding depression
in conservation biology, as a single value per population, deserves re-evaluation and further refinement.
Evidence is accumulating that inbreeding depression
may be a trait that is variable among families within a
population (e.g. Mutikainen & Delph 1998; Pico et al.
2004b; Bailey & McCauley 2006). A case study on
Scabiosa columbaria (Van Treuren et al. 1993; Pico et al.
2004b) explained the difference between a population
perspective and a genotype (cf. family) perspective on
inbreeding depression. Scabiosa columbaria has been
used as a model to investigate the effects of small population size on genetic variation within populations
(Ouborg et al. 1991; Van Treuren et al. 1991) and the
corresponding average fitness of individuals (Ouborg
1993a; Van Treuren et al. 1993). Inbreeding depression
as expressed in germination, seedling survival, juvenile
survival, adult survival, flowering percentage and
reproductive output was estimated for a large population of this species in the Netherlands. The performance
of selfed offspring compared with outcrossed offspring
ranged from 1.08 (i.e. inbreeding enhancement; juvenile
survival) to 0.65 (i.e. inbreeding depression; adult
survival) (Van Treuren et al. 1993). In 2003, the same
population was re-sampled, with a design that specifically aimed to quantify variance in inbreeding depression
among families. Although the average inbreeding
depression values in this second study were very similar
to those in the first study, high variance in inbreeding
depression was found among families for all life-history
traits (Pico et al. 2004b). Some families showed strong
inbreeding depression, as was expected based on the
findings of the first study. At the same time, however,
approximately as many families showed significant inbreeding enhancement, meaning that inbred individuals
outperformed outcrossed individuals. A third category
of families showed no inbreeding depression at all.
While the findings of the first study would lead to
the conclusion that inbreeding depression could be a
serious problem to this population, the second study
would result in the conclusion that inbreeding depression would only be a temporary threat. Over time,
selection would lead to an increase in the frequency of
families that show no inbreeding depression or even
benefit from inbreeding.
A further refinement of the manner in which inbreeding depression is incorporated in the conservation
genetics paradigm is stimulated by the observation that
inbreeding depression is variable between stages
(Husband & Schemske 1996). The study by Pico et al.
(2004b) found no correlation between the inbreeding
depression values that were found for different lifehistory stages. Furthermore, some evidence suggests
that inbreeding depression is not a linear function of
inbreeding level (Pray & Goodnight 1995; Ouborg
et al. 2000), which further complicates matters.
The origin of the observed variation in inbreeding
depression is not yet completely understood. Models
have been constructed to investigate the role of differences in inbreeding history, of mutation and of genetic
variation per se (Schultz & Willis 1995; Fowler & Whitlock 1999; Kelly 2004; Fox 2005; Moorad & Wade 2005).
Thus, although considerable theoretical attention has
been given to variance in inbreeding depression and its
consequences for the purging of genetic load (Whitlock
2000; Glemin 2003), there is a great need for empirical
studies that try to quantify this variance in the context
of the conservation genetics paradigm.
In summary, inbreeding depression values not only
seem to be dependent on population size, but seem to
vary between life-history stages, families and inbreeding
levels. This calls for a more sophisticated form of incorporation of inbreeding depression in conservation
genetics. We need to know the extent of variation in
inbreeding depression in each particular situation.
Given the dependence of extinction probabilities on
stochastic variation, the research agenda of conservation
genetics should include studies searching for the effects
of population size on the variance rather than the mean
of inbreeding depression. Besides being important for
the evaluation of extinction risks, variance in inbreeding
1240
N. J. Ouborg,
P. Vergeer &
C. Mix
depression might also play an important role in a
regional population or metapopulation context. The
future dynamics and genetics of a newly founded
population, and of the metapopulation as a whole, might
be very different when a site is colonized by a genotype
that experiences inbreeding enhancement rather than
inbreeding depression. There is a need for field studies
that measure the development of fitness over several
generations in experimental populations started with
different genotypes, in order to evaluate the impact of
variance in inbreeding depression on the probability of
extinction.
Neutral markers vs. functional genes
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
As explained in the section on the conservation genetics
paradigm, the effects of loss of genetic diversity are
two-fold: there are the immediate effects on fitness (e.g.
inbreeding depression) discussed in the previous section
(Fig. 1B), and there is the postponed effect on extinction
probability via the loss of evolutionary potential
(Fig. 1C). The effect of small population size on genetic
diversity and its loss has traditionally been investigated
using genetic markers. From the early 1980s onwards,
allozyme markers were used, but DNA markers, such
as microsatellites, AFLPs and ISSR markers, have
since replaced them, mainly because they are easy to
use when scoring many markers for many individuals
and because of their higher level of resolution compared
with allozymes.
Whether an observed reduction of genetic marker
variation in a small population necessarily indicates a
loss of evolutionary potential is debatable. Although
markers are excellent tools for studying the effects of
non-selective processes such as drift and inbreeding,
they are not necessarily suitable to quantify changes
in selectively important variation. In general, what
researchers did in the past was to measure neutral
variation with many markers, and assume that selectively
important markers would follow the same pattern (e.g.
O’Brien et al. 1985; Allendorf & Leary 1986). The
correlation between neutral marker variation and the
variation in quantitative traits that are important for
fitness is, however, not undisputed, with Lynch (1996)
arguing that this correlation is likely to be low. Some
meta-analysis reviews have indeed illustrated this weak
or absent correlation (Butlin & Tregenza 1998; Reed &
Frankham 2001). In a meta-analysis of 34 studies, Reed
& Frankham (2003) found a significant correlation
between fitness and genetic diversity in molecular
markers, but genetic diversity measured in this way
could only explain 19% of the total variation in fitness.
It is thus dangerous to draw conclusions about the
viability (and potential for evolutionary adaptation) of
populations exclusively based on observed levels of
molecular marker variation, and such conclusions are
better avoided. In general, our understanding of the
role of selection in small populations is limited, and
requires more theoretical and empirical research.
Whatever the outcome of the interplay between drift
and inbreeding on the one hand and selection on the
other, very few tools were until recently available to get
an empirical grip on selectively important genetic
variation. With the recent emergence of the field of ecogenomics, this situation is changing rapidly, as several
techniques from the genomics toolbox will allow
detailed studies of non-neutral genetic variation in
populations (Van Tienderen et al. 2002).
The screening of sequence variation in functional
genes is developing rapidly, as a result of the availability
of complete genome data of model species, the most
prominent of which is Arabidopsis thaliana. Although
each new species will require some work to fine-tune
the Arabidopsis technique for the species of interest,
this is becoming more and more common practice
(Jackson et al. 2002; Stearns & Magwene 2003). There
have already been many other examples besides Arabidopsis of the use of PCR-based methods to study variation in genes.
The genes involved in flowering in plants form a
useful example for illustrating the potential of an ecogenomic approach. The regulatory gene pathways
connecting environmental cues (light, day length and
temperature) to the timing of flowering and the
development of floral structures have been described in
considerable detail (Blazquez 2000). Although these
pathways are based on studies of A. thaliana, homologues of many genes in these pathways have been
identified in other plant species. For instance, the
LEAFY gene, which controls the identity of meristems
that may develop into flowers, has been sequenced in
species from the Leguminosae family (Archambault &
Bruneau 2004) and the Poaceae family (Bomblies &
Doebley 2005), as well as in Eucalyptus species (Southerton et al. 1998), apple (Wada et al. 2002), maize
(Bomblies et al. 2003), orchids (Montieri et al. 2004)
and Pinus (Mouradov et al. 1998).
Flowering time is controlled by two interacting
genes from this pathway, FRIGIDA (FRI) and FLOWERING LOCUS C (FLC). A large number of variants
of the FRI gene were detected when sequencing field
ecotypes of A. thaliana (Le Corre et al. 2002). Haplotypes were categorized by the presence or absence of
major deletions in the gene. A latitudinal cline in flowering time was found in Northern European and North
American ecotypes, and proved to be correlated with
variations in the FRI haplotype group, which lack the
major deletions and are therefore fully functional
(Stinchcombe et al. 2004). Variation within and among
12 French populations of A. thaliana was surveyed and
compared with variation at neutral microsatellite loci
(Le Corre 2005). A striking result was that the differentiation among populations in FRI was significantly
higher than that in microsatellites, suggesting that local
selection is affecting FRI variation.
While FRI and flowering time are among the more
fully researched examples, many other traits are about
to be investigated in the same way (Shimizu & Purugganan
1241
The rough edges of
the conservation
genetics paradigm
2005). This opens up the exciting possibility of investigating how sequence variation in functional genes subjected to selection is correlated with population size
and population isolation. It opens up new avenues to
improve our insight into the role of selection within the
context of the conservation genetics paradigm.
Another promising technique is the study of gene
expression, by means of real time PCR or microarrays
(Gibson 2002; Thomas & Klaper 2004). Although the
latter method will at this point in time have to build on
the availability of microarrays from well-studied species
(A. thaliana, Medicago truncatula, maize, rice, etc.),
the number of labs developing microarrays for their
own species is rapidly increasing (e.g. Hegarty et al.
2005).
An example of the gene expression approach is given
by Lai et al. (2006), who performed microarray analysis
of hybrid sunflower species. They screened the expression of around 3000 genes, and discovered that 12.8%
of them showed significant differences in expression
between the two parental species Helianthius annuus
and H. petiolaris, and the hybrid H. deserticola. A
growth study in the environment of H. deserticola
evaluated the adaptive significance of differential
expression of five of these genes. A significant association with fitness was found for one gene. Another
example is a recent field study by Schmidt & Baldwin
(2006), who, among other things, evaluated the transcriptional responses of Solanum nigrum in a field study
with competition as factor. They demonstrated, not
unexpectedly, that competition reduced the reproduction
of individual plants. At the same time they analysed the
expression differences, using a microarray with 568
known genes, between competitive and control plants,
and identified many genes that were either up- or downregulated in the competitive plants. These examples
show that gene expression studies can provide insights
into the adaptive potential of populations (Whitehead
& Crawford 2006).
In conclusion, the field of eco-genomics offers great
opportunities for conservation genetics. First, it offers
the tools to investigate the relationship between population size and population isolation and the extent of
adaptively important genetic variation. Second, it also
offers approaches to investigate in detail the relative
roles of genetic drift and selection within the context of
the conservation genetics paradigm. To profit fully
from these possibilities, it is required that for a number
of species, which could serve as conservation genetic
model species, a sequence programme is started and an
expressed sequence tag (EST) library is made and used
to design microarrays.
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
From single-species to multi-species approaches
The conservation genetics paradigm implicitly focuses
on single species. Although environmental impact on
genetics and demography are of course not ignored, the
problem is basically treated as involving a single species
(or population) in an influencing but non-responsive
world. That is, whereas the environment does affect the
demography and genetics of the species, the species
does not affect its environment. This is equally true for
abiotic and biotic environments.
Obviously, the functioning of a plant population in its
natural environment involves complex real interactions.
Plants interact with their abiotic (e.g. nutrients, moisture
level) and biotic environment (e.g. pollinators, herbivores,
pathogens, mycorhiza). Neither the abiotic nor the
biotic environment is an unchangeable, inert agent;
they are responding reagents (Fig. 1D).
Several studies have shown a decrease in pollinator
richness and abundance in response to habitat fragmentation (e.g. Kearns et al. 1998; Ashworth et al.
2004; Duncan et al. 2004). In recent years, studies of
the relationship between plant–pollinator interactions
and habitat fragmentation have been further extended
to include plant–herbivore as well as plant–pathogen
interactions (e.g. Ouborg et al. 2000; Carr & Eubanks
2002; Ivey et al. 2004). Tscharntke and co-workers have
demonstrated that habitat fragmentation may alter the
species web dramatically, by affecting different trophic
levels in different ways (e.g. Kruess & Tscharntke 1994,
2000; Zabel & Tscharntke 1998). The approach is characterized by a multi-species community approach that
until now, however, has largely ignored the genetic
effects of habitat fragmentation.
Effects of habitat fragmentation on community
structure may come about in several ways (Ouborg &
Biere 2003). First, fragmentation may directly affect
the presence and abundance of organisms at higher
trophic levels. For instance, smaller plant populations
offer smaller amounts of resources, which at some
threshold level become insufficient to sustain viable
herbivore populations (Anderson & May 1979; Getz &
Pickering 1983). Thus, small populations may be
protected to some extent from specialist herbivore and
disease pressure. Examples of this have been described
for natural plant–pathogen systems (Jennersten et al.
1983; Alexander 1990; Carlsson & Elmqvist 1992). An
example of a threshold effect in plant–herbivore interactions was found in Salvia pratensis populations in the
Netherlands. Salvia produces flowers which may each
produce up to four seeds. Seeds are infected by the gall
wasp Aylax salviae, which lays eggs in the developing
seed. It thereby induces the formation of seed galls, and
the reproductive output of an affected individual is
reduced, sometimes by as much as 60%. In a survey of
20 populations, only populations larger than 300
individuals were found to be infected (Fig. 4). (N.J.O.,
unpublished data).
More isolated populations may be less likely to
attract herbivores and diseases (Simberloff 1988).
Positive correlations have been found between disease
incidence in local populations and the disease status
of neighbouring populations (Burdon et al. 1995;
Antonovics et al. 1997; Ericson et al. 1999). While
increasing connectivity in a fragmented landscape has
continuous interaction with the focal species. Some
conclusions of conservation genetics might change
once the community structure is incorporated in a
dynamic approach. For instance, the general view that
small populations are at peril may be reconsidered to
some degree if we find that on the one hand small populations are protected against herbivores and diseases,
while on the other they may lose resistance alleles
owing to genetic erosion.
1242
N. J. Ouborg,
P. Vergeer &
C. Mix
Fig. 4 Relationship between population size (measured as
the number of flowering individuals) of Salvia pratensis
populations and the percentage of individuals infected by the
gall wasp Aylax salviae.
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
been used as a method to alleviate the effects of habitat
fragmentation on population viability (e.g. Harrison &
Fahrig 1995), increased connectivity could also result
in increased disease transmission. The net effect of
increased connectivity on population viability therefore depends on the balance between these positive and
negative effects (Hess 1996).
The second way in which habitat fragmentation may
affect community structure is via the effects of loss of
genetic diversity, which modifies the genotypic diversity in local populations. If the genotypic composition
of a plant population affects the dynamics of the organisms at the second trophic level, as is expected when
attractiveness, resistance, tolerance and/or avoidance
have a genetic basis, the loss of resistance and other
alleles from the plant population may change the entire
food web structure, including third, fourth, etc.,
trophic levels.
Third, genetic erosion effects may also be evident
from an increase in the average inbreeding level in small
populations. If inbreeding results in changes in the
physical, chemical and nutritional traits of individual
plants, it will alter the interaction between plants and
their pollinators, pathogens and herbivores. (Ajala
1992; Strauss & Karban 1994; Nunez-Farfán et al.
1996; Olff et al. 1999) As an indirect consequence, the
interaction between herbivores and their natural
enemies may be changed (e.g. Price et al. 1980; Hare
1992). Clear examples of this effect have been found in
studies investigating the effect of monocultures or low
biodiversity (e.g. Palmer & Maurer 1997; Knops 1999;
Dukes 2002).
Although these effects have been studied to some
extent, no coherent framework has yet been developed
to link the conservation genetics paradigm with community structure effects. Much more research should
be devoted to attempts to connect the genetic effects of
habitat fragmentation on plant populations with the
influence these effects may have on community structure.
A more dynamic perspective should be developed, in
which the biotic environment of the focal species is an
ecologically and evolutionarily responding agent, in
From phenomenology to general principles
Most empirical studies performed within the conservation genetics paradigm have concentrated on rare and
endangered species. Although the conservation of what
is threatened is in many cases the main and legitimate
motivator for the research, this may also introduce a
bias in the conclusions about the general applicability
of the results. Rare species occupy habitats that also
feature very common species. Even though the species
are all subjected to the same degree of landscape fragmentation, some species are strongly affected whereas
others do not respond at all. Exclusively studying rare
species may lead to results that are specific for the set of
traits that cause species to be unable to cope with fragmentation. If conservation genetics is to develop from
a phenomenological crisis discipline, which concentrates on case studies, into a multidisciplinary science
searching for the general principles that are involved
in determining extinction probabilities of plant populations, it is also necessary to study common species within
the context of the conservation genetics paradigm
(Fig. 1E).
There are three approaches available to use common
species to increase our understanding of the relationships that make up the conservation genetics paradigm,
as outlined in Fig. 1.
The first is a comparative approach, in which the
correlations in the diagram (Fig. 1) are determined for
common species and compared with the correlation
structure of rare or endangered species, which may
reveal which processes render a species endangered.
Some studies have indeed concentrated on common
species. Hooftman et al. (2004) studied common species
in fragmented Swiss fens and found some evidence that
genetic drift does indeed affect genetic diversity in
small populations of these species. A similar result was
obtained in a study on the effects of fragmentation on
the common plant species Succisa pratensis (Vergeer
et al. 2003). Jules (1998) investigated the effects of fragmentation on the demography of a common understory
herb of forests in the western USA, Trillium ovatum. He
found evidence that fragmentation would lead to a
decline in population persistence even in a common
species. By contrast, a study of the conservation genetics
of the co-dominant grass species Andropogon gerardii
and Sorghastrum nutans from North American prairies
by Gustafson et al. (2004a) found no evidence for a
relationship between either population size or population
1243
The rough edges of
the conservation
genetics paradigm
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
density and genetic diversity. However, a second study
by the same researchers (Gustafson et al. 2004b) did
find evidence for the effect of local adaptation on the
competitive ability of A. gerardii.
Thus, common species have received some attention
within the context of the conservation genetics
paradigm, including well-known work on Eichhornia
paniculata (e.g. Husband & Barrett 1992) and Lythrum
salicaria (e.g. Eckert et al. 1996). However, few studies
have deliberately used a comparative approach. Karron
(1989) compared inbreeding depression between rare
and widespread species of Astragalus. Cole (2003)
attempted to find general patterns in a comparison of
genetic variation in a total of 247 rare vs. common
plant species. He concluded that rare plant species
show more significant reductions in genetic variation
than common species. Although the results of that
comparison were most likely confounded by the very
different design of studies on common species as compared with rare species, they underline the importance
of incorporating research on common species in a
conservation biology context.
Second, the manipulation of common species offers
a way of examining components of the conservation
genetics paradigm in detail. Common species could
be used as a model to investigate processes that are
difficult or impossible to investigate for rare species,
because of their susceptibility and low numbers. This
approach has been used for several species, including
Arabidopsis thaliana (e.g. Cahill et al. 2005), Brassica
rapa (e.g. Wise et al. 2002) and Clarkia pulchella (Newman
& Pilson 1997).
A third promising approach is the use of invasive
plant species for conservation genetics research. These
species invade new areas, at which stage they must
temporarily be subjected to the same small and isolated
population effects. In other words, they go through a
bottleneck, and significant genetic drift effects can be
expected. Yet, invasive species are apparently fully
capable of dealing with these effects and developing
into thriving species. Petit et al. (2004) adopted this
approach for plants, by comparing species of oak
(Quercus sp.) to elucidate the difference between
invasive and non-invasive tree species. Other examples
include work on Butomus umbellatus (Kliber & Eckert
2005), Capsella bursa-pastoris (Neuffer & Hurka
1999), Alliaria petiolata (Meekins et al. 2001) and
introduced populations of Epipactis helleborine (Squirrell et al. 2001). The focus of most of these studies was
on understanding the causes of the invasiveness of the
species. Yet, studies of invasive species could provide
reference data for conservation genetics studies of rare
and endangered, conspecific, species. This comparative
approach may offer a road towards understanding the
general principles in conservation genetics.
In conclusion, the comparative study of rare,
common and invasive species will allow evaluation of
the conservation genetics paradigm, beyond the possibilities of case studies on endangered species alone.
Conclusions
The conservation genetics paradigm is currently widely
accepted in ecology, genetics and conservation biology.
With the paradigm becoming widely embraced, attention has shifted from the large and deterministic to the
small and stochastic. The paradigm has emphasized
the limits of theoretical models that implicitly assume
(near) infinite population sizes, it has focused attention
on the role of stochasticity in population dynamics and
population genetics, and it has been an invaluable
guideline for assessing the viability of endangered
species and ways of improving this.
This paper has tried to identify aspects of the paradigm (‘rough edges’) that could, or should, be refined.
Although conservation genetics research has provided
ample evidence for the importance of genetics in conservation, we have discussed five areas where theoretical
concepts and empirical approaches could be improved,
to achieve a more thorough understanding of the role
of genetics. These areas are also relevant to the ‘habitat
quality’ vs. ‘conservation genetics’ controversy. For
instance, although dispersal obviously has genetic
consequences, the rate of dispersal has ecological determinants; inbreeding depression, though clearly having
genetic causes, would be a meaningless concept without
referring to environmental conditions; species interactions may prove to be as important as local loss of
genetic diversity. These are not only areas where further
refinement of the conservation genetic paradigm can
be established, but research in these areas will also help
in the further reconciliation of the ‘habitat quality’ and
‘conservation genetics’ paradigms. Instead of viewing
ecology and genetics as opposite concepts, much can be
gained by assessing how ecology and genetics interact
in determining the viability of populations. The areas
discussed may define a research agenda to come to a more
complete understanding of this interaction between
ecology and genetics, and thereby a more thorough
understanding of the complex forces that determine
biodiversity.
Acknowledgements
The present paper is the reflection of many years of
research in the field of conservation biology. Although
the views are entirely the responsibility of the authors,
their development has benefited from discussions with
numerous people. We would specifically like to thank
Arjen Biere, Merel Soons, Eelke Jongejans, Ove Erikson,
Mike Hutchings, Diethard Matthies, Tomas Herben
and Jan van Groenendael. We would also like to thank
David Gibson and two anonymous referees for their
valuable comments on an earlier version of this paper.
N.J.O. was supported by the EU TRANSPLANT grant
(EVK2-1999-00042), P.V. was supported by a grant
from the Dutch Technology Foundation (STW) and
C.M. was supported by the Netherlands Organization
for Scientific Research (NWO project 805-33-451).
1244
N. J. Ouborg,
P. Vergeer &
C. Mix
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
References
Ajala, S.O. (1992) Inheritance of resistance in Maize to the
spotted stem-borer Chilo partellus (Swinhoe). Maydica, 37,
363 –359.
Alexander, H.M. (1990) Dynamics of plant–pathogen interactions in natural plant communities. Pests, Pathogens
and Plant Communities (eds J.J. Burdon & S.R. Leather),
pp. 31– 45. Blackwell, Oxford.
Allendorf, F.W. & Leary, R.F. (1986) Heterozygosity and
fitness in natural populations of animals. Conservation
Biology: the Science of Scarcity and Diversity (ed. M.E.
Soulé), pp. 57–76. Sinauer, Sunderland, MA.
Anderson, R.M. & May, R.M. (1979) Population biology of
infectious diseases I. Nature, 280, 361–367.
Antonovics, J., Thrall, P.H. & Jarosz, A.M. (1997) Genetics
and spatial ecology of species interactions: the Silene–
Ustilago. Spatial Ecology: the Role of Space in Population
Dynamics and Interspecific Interactions (eds D. Tilman &
P. Kareiva), pp. 158 –184. Princeton University Press,
Princeton, NJ.
Archambault, A. & Bruneau, A. (2004) Phylogenetic utility of
the LEAFY/FLORICAULA gene in the Ceasalpinioideae
(Leguminosae): gene duplication and a novel insertion.
Systematic Botany, 29, 609 –626.
Armbruster, P. & Reed, D.H. (2005) Inbreeding depression in
benign and stressful environments. Heredity, 95, 235 –242.
Ashworth, L., Aguilar, R., Galetto, L. & Aizen, M.A. (2004)
Why do pollination generalist and specialist plant species
show similar reproductive susceptibility to habitat fragmentation? Journal of Ecology, 92, 717–719.
Bailey, M.F. & McCauley, D.E. (2006) The effect of inbreeding,
outbreeding and long-distance gene-flow on survivorship
in North-American populations of Silene vulgaris. Journal
of Ecology, 94, 98 –109.
Barrett, S.C.H. & Kohn, J.R. (1991) Genetic and evolutionary
consequences of small population size in plants: implications for conservation. Genetics and Conservation of Rare
Plants (eds D.A. Falk & K.E. Holsinger), pp. 3 –30. Oxford
University Press, Oxford.
Bender, D.J., Contreras, T.H. & Fahrig, L. (1998) Habitat loss
and population decline: a meta-analysis of the patch size
effect. Ecology, 79, 517–533.
Bijlsma, R. & Loeschcke, V. (2005) Environmental stress,
adaptation and evolution: an overview. Journal of Evolutionary Biology, 18, 744 –749.
Blazquez, M.A. (2000) Flower developing pathways. Journal
of Cell Science, 113, 3547–3548.
Bomblies, K. & Doebley, J.F. (2005) Molecular evolution of
FLORICAULA/LEAFY orthologues in the Andropogoneae
(Poaceae). Molecular Biology and Evolution, 22, 1082 –1094.
Bomblies, K., Wang, R.-L., Ambrose, B.A., Schmidt, R.J.,
Meeley, R.B. & Doebley, J. (2003) Duplicate FORICAULA/
LEAFY homologs zfl1 and zfl2 control inflorescence
architecture and flower patterning in maize. Development,
130, 2385 –2395.
Booth, R.E. & Grime, J.P. (2003) Effects of genetic impoverishment on plant community diversity. Journal of Ecology,
91, 721–730.
Booy, G., Hendriks, R.J.J., Smulders, M.J.M., van Groenendael,
J.M. & Vosman, B. (2000) Genetic diversity and the survival
of populations. Plant Biology, 2, 379 –395.
Bouchy, P., Theodorou, K. & Couvet, D. (2006) Metapopulation viability: influence of migration. Conservation Genetics,
6, 75 – 85.
Boyce, M.S., Haridas, C.V. & Lee, C.T. (2006) Demography
in an increasingly variable world. Trends in Ecology and
Evolution, 21, 141–148.
Brook, B.W., Tonkyn, D.W., O’Grady, J.J. & Frankham, R.
(2002) Contribution of inbreeding to extinction risk in
threatened species. Conservation Ecology, 6, art. no. 16.
Brown, J.H. & Kodric-Brown, A. (1977) Turnover rates in
insular biogeography: effect of migration and extinction.
Ecology, 58, 445 – 449.
Brys, R., Jacquemyn, H., Endels, P., Van Rossum, F., Hermy,
M., Triest, L., De Bruyn, L. & Blust, D.E. (2004) Reduced
reproductive success in small populations of the selfincompatible Primula vulgaris. Journal of Ecology, 92, 5–
14.
Bullock, J.M. & Hodder, K.H. (1997) Reintroductions:
challenges and lessons for basic ecology. Trends in Ecology
and Evolution, 12, 68 – 69.
Burdon, J.J., Ericson, L. & Müller, W.J. (1995) Temporal and
spatial changes in a metapopulation of the rust pathogen
Triphragmium ulmariae and its host, Filipendula ulmaria.
Journal of Ecology, 83, 979 – 989.
Butlin, R.K. & Tregenza, T. (1998) Levels of genetic polymorphism: marker loci versus quantitative traits. Philosophical
Transactions of the Royal Society of London B, 353, 187–
198.
Cahill, J.F., Kembel, S.W. & Gustafson, D.J. (2005) Differential
genetic influences on competitive effect and response in
Arabidopsis thaliana. Journal of Ecology, 93, 958–967.
Cain, M.L., Milligan, B.G. & Strand, A.E. (2000) Long
distance seed dispersal in plant populations. American Journal
of Botany, 87, 1217–1227.
Carlsson, U. & Elmqvist, T. (1992) Epidemiology of anther-smut
disease (Microbotryum violaceum) and numeric regulation
of populations of Silene dioica. Oecologia, 90, 509–517.
Carr, D.E. & Dudash, M.R. (2003) Recent approaches into
the genetic basis of inbreeding depression in plants.
Philosophical Transactions of the Royal Society of London
B, 358, 1071–1084.
Carr, D.E. & Eubanks, M. (2002) Inbreeding alters resistance
to insect herbivory and host plant quality in Mimulus
guttatus (Scrophulariaceae). Evolution, 56, 22 –30.
Charlesworth, D. & Charlesworth, B. (1987) Inbreeding
depression and its evolutionary consequences. Annual
Review of Ecology and Systematics, 18, 237–268.
Charlesworth, B. & Charlesworth, D. (1999) The genetic basis
of inbreeding. Genetics Research Cambridge, 74, 329–340.
Cody, M.L. & Overton, J.M. (1996) Short-term evolution of
reduced dispersal in island plant populations. Journal of
Ecology, 84, 53–61.
Colas, B., Olivieri, I. & Riba, M. (1997) Centaurea corymbosa,
a cliff-dwelling species tottering on the brink of extinction:
a demographic and genetic study. Proceedings of the
National Academy of Sciences of the USA, 94, 3471–3476.
Cole, C.T. (2003) Genetic variation in rare and common
plants. Annual Review of Ecology and Systematics, 34, 213–
237.
Crnokrak, P. & Barrett, S.C.H. (2002) Perspective: purging
the genetic load: a review of the experimental evidence.
Evolution, 56, 2347–2358.
De Kroon, H., Plaisier, A., Van Groenendael, J. & Caswell, H.
(1986) Elasticity – the relative contribution of demographic
parameters to population growth rate. Ecology, 67, 1427–
1431.
Dennis, B. (2002) Allee effects in stochastic populations.
Oikos, 96, 389–401.
Donohue, K. (1999) Seed dispersal as a maternally influenced
character: mechanistic basis of maternal effects and selection on maternal characters in an annual plant. American
Naturalist, 154, 674–689.
Drake, J.M. (2005) Density-dependent demographic variation
determines extinction rate of experimental populations.
PLOS Biology, 3, 1300–1304.
Dukes, J.S. (2002) Species composition and diversity affect
grassland susceptibility and response to invasion. Ecological
Applications, 12, 602–617.
Duncan, D.H., Nicotra, A.B., Wood, J.T. & Cunningham,
S.A. (2004) Plant isolation reduces outcross pollen receipt
1245
The rough edges of
the conservation
genetics paradigm
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
in a partially self-compatible herb. Journal of Ecology, 92,
977–985.
Dupré, C. & Ehrlén, J. (2002) Habitat configuration, species
traits and plant distributions. Journal of Ecology, 90, 796 –
805.
Eckert, C.G., Manicacci, D. & Barrett, S.C.H. (1996) Genetic
drift and founder effect in native versus introduced
populations of an invading plant, Lythrum salicaria
(Lythraceae). Evolution, 50, 1512–1519.
Ellstrand, N. & Elam, D.R. (1993) Population genetic consequences of small population size; implications for plant
conservation. Annual Review of Ecology and Systematics,
24, 217–242.
Engen, S., Bakke, O. & Islam, A. (1998) Demographic and
environmental stochasticity – concepts and definitions.
Biometrics, 54, 840 –846.
Ericson, L., Burdon, J.J. & Müller, W.J. (1999) Spatial and
temporal dynamics of epidemics of the rust fungus Uromyces
valerianae on populations of its host, Valeriana salina. Journal
of Ecology, 87, 649– 658.
Fagan, W.F. & Holmes, E.E. (2006) Quantifying the extinction
vortex. Ecology Letters, 9, 51–60.
Fahrig, L. (2003) Effects of habitat fragmentation on biodiversity. Annual Review of Ecology and Systematics, 34,
487–515.
Fowler, K. & Whitlock, M.C. (1999) The variance in inbreeding
depression and the recovery of fitness in bottlenecked
populations. Proceedings of the Royal Society of London B,
266, 2061–2066.
Fox, C.W. (2005) Problems in measuring among-family
variation in inbreeding depression. American Journal of
Botany, 92, 1929–1932.
Fox, G.A. & Kendall, B.E. (2002) Demographic stochasticity
and the variance reduction effect. Ecology, 83, 1928 –1934.
Frankel, O.H. & Soulé, M.E. (1981) Conservation and Evolution.
Cambridge University Press, Cambridge.
Frankham, R. (1995) Conservation genetics. Annual Review
of Genetics, 29, 305 –327.
Frankham, R., Ballou, J.D. & Briscoe, D.A. (2002) Introduction
to Conservation Genetics. Cambridge University Press,
Cambridge.
Frankham, R. & Ralls, K. (1998) Inbreeding leads to extinction.
Nature, 392, 441– 442.
Getz, W.M. & Pickering, J. (1983) Epidemic models: thresholds
and population regulation. American Naturalist, 121, 892–
898.
Gibson, G. (2002) Microarrays in ecology and evolution: a
preview. Molecular Ecology, 11, 17–24.
Glemin, S. (2003) How are deleterious mutations purged?
Drift versus non-random mating. Evolution, 57, 2678 –
2687.
Glemin, S., Ronfort, J. & Bataillou, T. (2003) Patterns of
inbreeding depression and the architecture of the load in
subdivided populations. Genetics, 165, 2193–2212.
Goodman, D. (1987) The demography of chance extinction.
Viable Populations for Conservation (ed. M.E. Soulé),
pp. 11–34. Cambridge University Press, Cambridge.
Gustafson, D.J., Gibson, D.J. & Nickrent, D.L. (2004a)
Conservation genetics of two-dominant grass species in an
endangered grassland ecosystem. Journal of Applied Ecology,
41, 389–397.
Gustafson, D.J., Gibson, D.J. & Nickrent, D.L. (2004b)
Competitive relationships of Andropogon gerardii (Big
Bluestem) from remnant and restored native populations
and select cultivated varieties. Functional Ecology, 18, 451–
457.
Hanski, I. & Gagiotti, O.E. (2004) Ecology, Genetics and Evolution of Metapopulations. Academic Press, Amsterdam.
Hare, J.D. (1992) Effects of plant variation on herbivore–
natural enemy interactions. Plant Resistance to Herbivores
and Pathogens: Ecology, Evolution and Genetics (eds R.S.
Fritz & E.L. Simms), pp. 278–298. The University of Chicago
Press, Chicago.
Harrison, S. & Fahrig, L. (1995) Landscape pattern and
population conservation. Mosaic Landscapes and Ecological
Processes (eds L. Hansson, L. Fahrig & G. Merriam),
pp. 293 –308. Chapman & Hall, London.
Hedrick, P.W. (2005a) Genetics of Populations, 3rd edn. Jones
and Bartlett, Boston.
Hedrick, P. (2005b) ‘Genetic restoration’: a more comprehensive
perspective than ‘genetic rescue’. Trends in Ecology and
Evolution, 20, 109.
Hedrick, P.W., Gadan, J. & Page, R.E. (2006) Genetic sex
determination and extinction. Trends in Ecology and
Evolution, 21, 55 –57.
Hegarty, M.J., Jones, J.M., Wilson, I.D., Barker, G.L.,
Coghill, J.A., Sanchez-Baracaldo, P., Liu, G., Buggs,
R.J.A., Abbott, R.J., Edwards, K.J. & Hiscock, S.J. (2005)
Development of anonymous cDNA microarrays to study
changes to the Senecio floral transcriptome during hybrid
speciation. Molecular Ecology, 14, 2493 –2510.
Hess, G. (1996) Disease in metapopulation models: implications for conservation. Ecology, 77, 1617–1632.
Higgins, K. & Lynch, M. (2001) Metapopulation extinction
caused by mutation accumulation. Proceedings of the
National Academy of Sciences of the USA, 98, 2928–2933.
Holsinger, K.E., Mason-Gamer, R.J. & Whitton, J. (1999)
Genes, demes, and plant conservation. Genetics and the
Extinction of Species (eds L.F. Landweber & A.P. Dobson),
pp. 23–46. Princeton University Press, Oxford.
Hooftman, D.A.P., Billeter, R.C., Schmid, B. & Diemer, M.
(2004) Genetic effects of habitat fragmentation on common
species of Swiss fen meadows. Conservation Biology, 18,
1043 –1051.
Hooftman, D.A.P. & Diemer, M. (2002) Effects of small
habitat size and isolation on the population structure of
common wetland species. Plant Biology, 4, 720–728.
Hufford, K.M. & Mazer, S.J. (2003) Plant ecotypes: genetic
differentiation in the age of ecological restoration. Trends in
Ecology and Evolution, 18, 147–155.
Husband, B.C. & Barrett, S.C.H. (1992) Effective population
size and genetic drift in tristylous Eichhornia paniculata
(Pontederiaceae). Evolution, 46, 1875 –1890.
Husband, B.C. & Schemske, D.W. (1996) Evolution and the
magnitude and timing of inbreeding depression in plants.
Evolution, 50, 54 –70.
Ingvarsson, P.K. (2001) Restoration of genetic variation lost –
the genetic rescue hypothesis. Trends in Ecology and Evolution,
16, 62–63.
Ivey, C.T., Carr, D.E. & Eubanks, M.D. (2004) Effects of
inbreeding in Mimulus guttatus on tolerance to herbivory in
natural environments. Ecology, 85, 567–574.
Jackson, R.B., Linder, C.R., Lynch, M., Purugganan, M.,
Somerville, S. & Thayer, S. (2002) Linking molecular
insight and ecological research. Trends in Ecology and
Evolution, 17, 409– 414.
Jennersten, O., Nilsson, S.G. & Wastljung, U. (1983) Local
plant populations as ecological islands: the infection of
Viscaria vulgaris by the fungus Ustilago violacea. Oikos, 41,
391–395.
Jules, E.S. (1998) Habitat fragmentation and demographic
change for a common plant: Trillium in old-growth forest.
Ecology, 79, 1645 –1656.
Karron, J.D. (1989) Breeding systems of inbreeding depression in geographically restricted and widespread species of
Astragalus (Fabaceae). American Journal of Botany, 76,
331–340.
Kearns, C.A., Inouye, D.W. & Waser, N.M. (1998) Endangered
mutualisms: the conservation of plant–pollinator interactions.
Annual Review of Ecology and Systematics, 29, 83–112.
Keller, L.F. & Waller, D.M. (2002) Inbreeding effects in wild
populations. Trends in Ecology and Evolution, 17, 230–241.
1246
N. J. Ouborg,
P. Vergeer &
C. Mix
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
Kelly, J.K. (2004) Family level inbreeding depression and the
evolution of plant mating systems. New Phytologist, 165,
55 – 62.
Kliber, A. & Eckert, C.G. (2005) Interaction between founder
effect and selection during biological invasion in an aquatic
plant. Evolution, 59, 1900 –1913.
Knops, J.M.H. (1999) Effects of plant species richness on
invasion dynamics, disease outbreaks, insect abundance
and diversity. Ecology Letters, 2, 286 –293.
Kolb, A. (2005) Reduced reproductive success and offspring
survival in fragmented populations of the forest herb
Phyteuma spicatum. Journal of Ecology, 93, 1226 –1237.
Krauss, J., Klein, A.M., Steffen-Dewenter, I. & Tscharntke, T.
(2004) Effects of habitat area, isolation, and landscape
diversity on plant species richness of calcareous grasslands.
Biodiversity and Conservation, 13, 1427–1439.
Krebs, C.J. (1972) Ecology. Harper & Row, New York.
Kruess, A. & Tscharntke, T. (1994) Habitat fragmentation,
species loss and biological control. Science, 264, 1581–
1584.
Kruess, A. & Tscharntke, T. (2000) Species richness and
parasitism in a fragmented landscape: experiments and
field studies with insects on Vicia sepium. Oecologia, 122,
129 –137.
Kuhn, T.S. (1996) The Structure of Scientific Revolutions, 3rd
edn. University of Chicago Press, Chicago.
Lai, Z., Gross, B.L., Zou, Y., Andrews, J. & Rieseberg, L.H.
(2006) Microarray analysis reveals differential gene expression in hybrid sunflower species. Molecular Ecology, 15,
1213 –1227.
Lande, R. (1988) Genetics and demography in biological
conservation. Science, 241, 1455 –1460.
Lande, R. (1993) Risks of population extinction from
demographic and environmental stochasticity and random
catastrophes. American Naturalist, 142, 911–927.
Le Corre, V. (2005) Variation at two flowering time genes
within and among populations of Arabidopsis thaliana:
comparison with markers and traits. Molecular Ecology,
14, 4181– 4192.
Le Corre, V., Roux, F. & Reboud, X. (2002) DNA polymorphism
at the FRIGIDA gene in Arabidopsis thaliana: extensive
nonsynonimous variation is consistent with local selection
for flowering time. Molecular Biology and Evolution, 19,
1261–1271.
Leimar, O. & Norberg, U. (1997) Metapopulation extinction
and genetic variation in dispersal related traits. Oikos, 80,
448–458.
Leimu, R., Mutikainen, P., Koricheva, J. & Fischer, M. (2006)
How general are positive relationships between plant
population size, fitness, and genetic variation? Journal of
Ecology, 94, 942–952.
Lienert, J. & Fischer, M. (2003) Habitat fragmentation affects
the common wetland specialist Primula farinose in northeast Switzerland. Journal of Ecology, 91, 587–599.
Linhart, Y.B. & Grant, M.C. (1996) Evolutionary significance
of local genetic differentiation in plants. Annual Review of
Ecology and Systematics, 27, 237–277.
Luijten, S.H., Kéry, M., Oostermeijer, J.G.B. & Den Nijs,
H.C.M. (2002) Demographic consequences of inbreeding
and outbreeding in Arnica Montana: a field experiment.
Journal of Ecology, 90, 593– 603.
Lynch, M. (1996) A quantitative-genetic perspective on conservation issues. Conservation Genetics: Case Histories
from Nature (eds J.C. Avise & J.L. Hamrick), pp. 471–501.
Chapman and Hall, New York.
Lynch, M., Connery, J. & Burger, R. (1995) Mutation
accumulation and the extinction of small populations.
American Naturalist, 146, 489 –518.
Matthies, D., Brauer, I., Maibom, W. & Tscharntke, T. (2004)
Population size and the risk of local extinction: empirical
evidence from rare plants. Oikos, 105, 481– 488.
McCarthy, M.A., Thompson, C.J. & Possingham, H.P. (2005)
Theory for designing nature reserves for single species.
American Naturalist, 165, 250 –257.
McPeek, M.H. & Holt, R.D. (1992) The evolution of dispersal
in spatially and temporally varying environments. American Naturalist, 140, 1010 –1027.
Meekins, J.F., Ballard, H.E. & McCarthy, B.C. (2001) Genetic
variation and molecular biogeography of a North American
invasive plant species (Alliaria petiolata, Brassicaceae).
International Journal of Plant Science, 162, 161–169.
Mills, L.S. & Smouse, P.E. (1994) Demographic consequences
of inbreeding in remnant populations. American Naturalist, 144, 412– 431.
Mitton, J.B. (1997) Selection in Natural Populations. Oxford
University Press, Oxford.
Mitton, J.B. & Grant, M.C. (1984) Associations among
protein heterozygosity, growth rate, and developmental
homeostasis. Annual Review of Ecology and Systematics,
15, 479– 499.
Mix, C., Pico, F.X., van Groenendael, J.M. & Ouborg, N.J.
(2006) Inbreeding and soil conditions affect dispersal and
components of performance of two plant species in fragmented landscapes. Basic and Applied Ecology, 7, 59–69.
Moilanen, A. & Nieminen, M. (2002) Simple connectivity
measures in spatial ecology. Ecology, 83, 1131–1145.
Montieri, S., Gaudio, L. & Aceto, S. (2004) Isolation of the
LFY/FLO homologue in Orchis italica and evolutionary
analysis in some European orchids. Gene, 333, 101–109.
Moorad, J.A. & Wade, M.J. (2005) A genetic interpretation of
the variation in inbreeding depression. Genetics, 170, 1373–
1384.
Mouradov, A., Glassick, T., Hamdorf, B., Murphy, L.,
Fowler, B., Marla, S. & Teasdale, R.D. (1998) NEEDLY, a
Pinus radiate ortholog of FORICAULA/LEAFY genes,
expressed in both reproductive and vegetative meristems.
Proceedings of the National Academy of Sciences of the
USA, 95, 6537– 6542.
Mutikainen, P. & Delph, L.F. (1998) Inbreeding depression in
gynodioecious Lobelia siphilitica: among-family differences
override between-morph differences. Evolution, 52, 1572–
1582.
Neuffer, B. & Hurka, H. (1999) Colinization history and
introduction dynamics of Capsella bursa-pastoris (Brassicaceae) in North America: isozymes and quantitative traits.
Molecular Ecology, 8, 1667–1681.
Newman, D. & Pilson, D. (1997) Increased probability of
extinction due to decreased genetic effective population size:
experimental populations of Clarkia pulchella. Evolution,
51, 354 –362.
Nicolé, F., Brzosko, E. & Till-Bottraud, I. (2005) Population
viability analysis of Cypripedium calceolus in a protected
area: longevity, stability and persistence. Journal of Ecology,
93, 716 –726.
Nunez-Farfán, J., Cabralesvargas, R.A. & Dirzo, R. (1996)
Mating system consequences on resistance to herbivory
and life history traits in Datura stramonium. American
Journal of Botany, 83, 1041–1049.
Nunney, L. & Campbell, K.A. (1993) Assessing minimum
viable population size: demography meets population
genetics. Trends in Ecology and Evolution, 8, 234–239.
O’Brien, S.J., Roelke, M.E., Marker, L., Newman, A., Winkler,
C.A., Meltzer, D., Colly, L., Evermann, J.F., Bush, M. &
Wildt, D.E. (1985) Genetic basis for species vulnerability in
the cheetah. Science, 227, 1428 –1434.
Olff, H., Brown, V.K. & Drent, R.H. (1999) Herbivores:
Between Plants and Predators. Blackwell, Oxford.
Olivieri, I. & Gouyon, P. (1997) Evolution of migration rates
and other traits: the metapopulation effect. Metapopulation
Biology: Ecology, Genetics and Evolution (eds I. Hanski &
M.E. Gilpin), pp. 293–323. Academic Press, San Diego.
Olivieri, I., Michalakis, Y. & Gouyon, P. (1995) Metapopulation
1247
The rough edges of
the conservation
genetics paradigm
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
genetics and the evolution of dispersal. American Naturalist,
146, 202 –227.
Oostermeijer, J.G.B., Luijten, S.H. & Den Nijs, J.C.M. (2003)
Integrating demographic and genetic approaches in plant
conservation. Biological Conservation, 113, 389 –398.
Ouborg, N.J. (1993a) On the relative contribution of genetic
erosion to the chance of population extinction. PhD thesis,
University of Utrecht, The Netherlands.
Ouborg, N.J. & Biere, A. (2003) The relationship between
plant-pathogen and plant–herbivore interactions and plant
population persistence in a fragmented landscape. Population
Viability Analysis in Plants (eds C.A. Brigham & M.W.
Schwartz), pp. 99–116. Springer-Verlag, Berlin.
Ouborg, N.J., Biere, A. & Mudde, C.L. (2000) Inbreeding
effects on biochemical resistance and transmission-related
traits in the Silene-Microbotryum pathosystem. Ecology,
81, 520 –531.
Ouborg, N.J. & Eriksson, O. (2004) Toward a metapopulation
concept for plants. Ecology, Genetics, and Evolution of
Metapopulations (eds I. Hanski & O.E. Gaggiotti), pp. 447–
469. Elsevier, San Diego.
Ouborg, N.J., Piquot, Y. & Van Groenendael, J.M. (1999)
Population genetics, molecular markers and the study of
dispersal in plants. Journal of Ecology, 87, 551–568.
Ouborg, N.J., van Treuren, R. & van Damme, J.M.M. (1991)
The significance of genetic erosion in the process of extinction
II. Morphological variation and fitness components in
populations of varying size of Salvia pratensis L. and Scabiosa
columbaria L. Oecologia, 86, 359–367.
Ozinga, W.A., Schaminee, J.H.J., Bekker, R.M., Bonn, S.,
Poschlod, P., Tackenberg, O., Bakker, J. & Van Groenendael,
J.M. (2005) Predictability of plant species composition
from environmental conditions is constrained by dispersal
limitation. Oikos, 108, 555–561.
Palmer, M.W. & Maurer, T.A. (1997) Does diversity beget
diversity? A case study of crops and weeds. Journal of
Vegetative Science, 8, 235–240.
Petit, R.J., Bialozyt, R., Garnier-Gére, P. & Hampe, A. (2004)
Ecology and genetics of tree invasions: from recent introductions to quaternary migrations. Forest Ecology and
Management, 197, 117–137.
Pico, F.X., Ouborg, N.J. & van Groenendael, J.M. (2004a)
Influence of selfing and maternal effects on life-cycle traits
and dispersal ability in the herb Hypochaeris radicata
(Asteraceae). Botanical Journal of the Linnean Society, 146,
383 –394.
Pico, F.X., Ouborg, N.J. & van Groenendael, J.M. (2004b)
Evaluation of the extent of among-family variation in
inbreeding depression in the perennial herb Scabiosa
columbaria (Dipsacaceae). American Journal of Botany, 91,
1183 –1189.
Pray, L.A. & Goodnight, C.J. (1995) Genetic variation in
inbreeding depression on the red flour beetle Tribolium
castaneum. Evolution, 49, 176 –188.
Prentice, H.C., Lönn, M., Rosquist, G., Ihse, M. & Kindström,
M. (2006) Gene diversity in a fragmented population of
Briza media: grassland continuity in a landscape context.
Journal of Ecology, 94, 87–97.
Price, P.W., Bouton, C.E., Gross, P., McPheron, B.A.,
Thompson, B.A. & Weiss, A.E. (1980) Interactions among
three trophic levels: influence of plants on interactions
between insect herbivores and natural enemies. Annual
Review of Ecology and Systematics, 11, 41– 65.
Reed, D.H. & Frankham, R. (2001) How closely correlated
are molecular and quantitative measures of genetic variation?
A meta-analysis. Evolution, 55, 1095 –1103.
Reed, D.H. & Frankham, R. (2003) Correlation between fitness
and genetic diversity. Conservation Biology, 17, 230 –237.
Richards, C.M. (2000) Inbreeding depression and genetic
rescue in a plant metapopulation. American Naturalist, 155,
383 –394.
Ronce, O., Bracket, S., Olivieri, I., Gouyon, P.H. & Clobert, J.
(2005) Plastic changes in seed dispersal along ecological
succession: theoretical predictions from an evolutionary
model. Journal of Ecology, 93, 431– 440.
Ronce, O., Perret, F. & Olivieri, I. (2000) Landscape dynamics
and evolution of colonizer syndromes: interactions between
reproductive effort and dispersal in a metapopulation.
Evolutionary Ecology, 14, 233 –260.
Schemske, D.W., Husband, B.C., Ruckelshaus, M.H., Goodwillie, C., Parker, I.M. & Bishop, J.G. (1994) Evaluating
approaches to the conservation of rare and endangered
plants. Ecology, 75, 584 – 606.
Schmidt, D.D. & Baldwin, I.T. (2006) Transcriptional
responses of Solanum nigrum to methyl jasmonate and
competition: a glasshouse and field study. Functional Ecology,
20, 500–508.
Schmitt, J., Ehrhardt, C.W. & Schwartz, D. (1985) Differential
dispersal of self-fertilized and outcrossed progeny in
jewelweed (Impatiens capensis). American Naturalist, 126,
570 –575.
Schoen, D.J., David, J.L. & Bataillou, T.M. (1998) Deleterious
mutation accumulation and the regeneration of genetic
resources. Proceedings of the National Academy of Sciences
of the USA, 95, 394–399.
Schonewald-Cox, C.S., Chambers, S.M., MacBryde, B. &
Thomas, L. (1983) Genetics and Conservation: a Reference
for Managing Wild Animal and Plant Populations. BenjaminCummings, London.
Schultz, S.T. & Willis, J.H. (1995) Individual variation in
inbreeding depression: the roles of inbreeding history and
mutation. Genetics, 141, 1209 –1223.
Shimizu, K.K. & Purugganan, M.D. (2005) Evolutionary and
ecological genomics of Arabidopsis. Plant Physiology, 138,
578 –584.
Simberloff, D.S. (1988) The contribution of population and
community biology to conservation science. Annual Review
of Ecology and Systematics, 19, 473 –511.
Soons, M.B. & Heil, G.W. (2002) Reduced colonization
capacity in fragmented populations of wind-dispersed
grassland forbs. Journal of Ecology, 90, 1033–1043.
Soons, N.M., Heil, G.W., Nathan, R. & Katul, G.G. (2004)
Determinants of long-distance dispersal by wind in grasslands. Ecology, 85, 3056–3068.
Soons, M.B., Messelink, J.H., Jongejans, E. & Heil, G.W.
(2005) Habitat fragmentation reduces grassland connectivity
for both short-distance and long-distance wind-dispersed
forbs. Journal of Ecology, 93, 1214–1225.
Soulé, M.E. (1986) Conservation Biology: the Science of Scarcity
and Diversity. Sinauer Associates, Sunderland, MA.
Soulé, M.E. & Wilcox, B.A. (1980) Conservation Biology:
an Evolutionary–Ecological Perspective. Sinauer Associates,
Sunderland, MA
Southerton, S.G., Strauss, S.H., Olive, M.R., Harcourt, R.L.,
Decroocq, V., Zhu, X.M., Llewellyn, D.J., Peacock, W.J. &
Dennis, E.S. (1998) Eucalyptus has a functional equivalent
of the Arabidopsis floral meristem identity gene LEAFY.
Plant Molecular Biology, 37, 897–910.
Spielman, D., Brook, B.W. & Frankham, R. (2004) Most
species are not driven to extinction before genetic factors
impact them. Proceedings of the National Academy of Sciences of the USA, 101, 15261–15264.
Squirrell, J., Hollingsworth, P.M., Bateman, R.M., Dickson,
J.H., Light, M.H.S., MacConnaill, M. & Tebbitt, M.C.
(2001) Partitioning and diversity of nuclear and organelle
markers in native and introduced ppulations of Epipactis
helleborine (Orchidaceae). American Journal of Botany, 88,
1409 –1418.
Stearns, S.C. & Magwene, P. (2003) The naturalist in a world
of genomics. American Naturalist, 161, 171–180.
Stinchcombe, J.R., Weinig, C., Ungerer, M., Olsen, K.M.,
Mays, C., Halldorsdottir, S.S., Purugganan, M.D. & Schmitt,
1248
N. J. Ouborg,
P. Vergeer &
C. Mix
© 2006 The Authors
Journal compilation
© 2006 British
Ecological Society,
Journal of Ecology,
94, 1233–1248
J. (2004) A latitudinal cline in flowering time in Arabidopsis
thaliana modulated by the flowering time gene FRIGIDA.
Proceedings of the National Academy of Sciences of the
USA, 101, 4712 – 4717.
Strauss, S.Y. & Karban, R. (1994) The significance of outcrossing in an intimate plant-herbivore relationship 1. Does
outcrossing provide an escape from herbivores adapted to
the parent plant? Evolution, 48, 454–464.
Tackenberg, O. (2003) Modelling long-distance dispersal of
plant diasporas by wind. Ecological Monographs, 73, 173 –
189.
Tackenberg, O., Poschlod, P. & Bonn, S. (2003) Assessment of
wind dispersal potential in plant species. Ecological Monographs, 73, 191–205.
Tallmon, D.A., Luikart, G. & Waples, R.S. (2004) The alluring
simplicity and complex reality of genetic rescue. Trends in
Ecology and Evolution, 19, 489– 496.
Theodorou, K. & Couvet, D. (2006) Genetic load in subdivided populations: interactions between the migration
rate, the size and the number of subpopulations. Heredity,
96, 69 –78.
Thomas, M.A. & Klaper, R. (2004) Genomics for the
ecological toolbox. Trends in Ecology and Evolution, 19,
439– 445.
Travis, J.M.J. & Dytham, C. (1999) Habitat persistence, habitat
availability and the evolution of dispersal. Proceedings of
the Royal Society of London B, 226, 723 –728.
Van Groenendael, J.M., Ouborg, N.J. & Hendriks, R.H.H.
(1998) Criteria for the introduction of plant species. Acta
Botanica Neerlandica, 47, 3 – 13.
Van Tienderen, P.H., De Haan, A.A., Van der Linden, G.C. &
Vosman, B. (2002) Biodiversity assessment using markers
for ecologically important traits. Trends in Ecology and
Evolution, 17, 577– 582.
Van Treuren, R., Bijlsma, R., Ouborg, N.J. & Van Delden, W.
(1993) The significance of genetic erosion in the process of
extinction IV. Inbreeding depression and heterosis effects
caused by selfing and outcrossing in Scabiosa columbaria.
Evolution, 47, 1669–1680.
Van Treuren, R., Bijlsma, R., Van Delden, W. & Ouborg, N.J.
(1991) The significance of genetic erosion in the process of
extinction I. Genetic differentiation in Salvia pratensis
and Scabiosa columbaria in relation to population size.
Heredity, 66, 181–189.
Vergeer, P. (2005) Introduction of threatened species in a fragmented and deteriorated landscape. PhD thesis, Radboud
University Nijmegen, The Netherlands.
Vergeer, P., Rengelink, R., Copal, A. & Ouborg, N.J. (2003)
The interacting effects of genetic variation, habitat quality
and population size on performance of Succisa pratensis.
Journal of Ecology, 91, 18 –26.
Vergeer, P., Sonderen, E. & Ouborg, N.J. (2004) Introduction
strategies put to the test: local adaptation versus heterosis.
Conservation Biology, 18, 812– 821.
Wada, M., Cao, Q., Kotoda, N., Soejima, J. & Masuda, T.
(2002) Apple has two ortologues of FORICAULA/
LEAFY involved in flowering. Plant Molecular Biology, 49,
567–577.
Whitehead, A. & Crawford, D.L. (2006) Variation within and
between species in gene expression: raw material for evolution.
Molecular Ecology, 15, 1197–1211.
Whitlock, M.C. (2000) Fixation of new alleles and the extinction of small populations: drift load, beneficial alleles, and
sexual selection. Evolution, 54, 1855 –1866.
Wise, C.A., Ranker, T.A. & Linhart, Y.B. (2002) Modelling
problems in conservation genetics with Brassica rapa:
genetic variation and fitness in plants under mild, stable
conditions. Conservation Biology, 16, 1542–1554.
Wright, S. (1969) Evolution and the Genetics of Populations,
Vol. II. The Theory of Gene Frequencies. The University of
Chicago Press, Chicago.
Young, A.G., Boyle, T. & Brown, A.D.H. (1996) The population
genetic consequences of habitat fragmentation for plants.
Trends in Ecology and Evolution, 11, 413 – 418.
Young, A.G. & Clarke, G.M. (2000) Genetics, Demography
and Viability of Fragmented Populations. Cambridge University Press, Cambridge.
Zabel, J. & Tscharntke, T. (1998) Does fragmentation of
Urtica habitats affect phytophagous and predatory insects
differentially? Oecologia, 116, 419– 425.
Received 14 February 2006
Revision accepted 15 June 2006
Handling Editor: David Gibson