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Ecosystems (2005) 8: 657–667 DOI: 10.1007/s10021-003-0138-8

Ecological and Evolutionary Consequences of Biological Invasion and Habitat Fragmentation Thomas S. Hoffmeister,1* Louise EM. Vet,2 Arjen Biere,3 Kent Holsinger,4 and Juliane Filser5 1

Institute of Ecology and Evolutionary Biology (IFOE), University of Bremen, Leobener Str., 28359 Bremen, Germany; 2Centre for Terrestrial Ecology, Department of Multitrophic Interactions, Netherlands Institute of Ecology (NIOO-KNAW), P.O. Box 40, 6666 ZG, Heteren, The Netherlands; 3Centre for Terrestrial Ecology, Department of Plant Population Biology, Netherlands Institute of Ecology (NIOO-KNAW), P.O. Box 40, 6666 ZG, Heteren, The Netherlands; 4Department of Ecology and Evolutionary Biology, University of Connecticut, Storrs-Mansfield, Connecticut 06269, USA; 5Department of General and Theoretical Ecology, UFT, University of Bremen, Leobener Str., 28359 Bremen, Germany

ABSTRACT There is substantial evidence that environmental changes on a landscape level can have dramatic consequences for the species richness and structure of food webs as well as on trophic interactions within such food webs. Thus far, the consequences of environmental change, and particularly the effects of invasive species and the fragmentation and isolation of natural habitats, have most often been studied in a purely ecological context, with the main emphasis on the description of alterations in species abundance and diversity and trophic links within food webs. Here, we argue that the study of evolutionary processes that may be affected by such changes is urgently needed to enhance our understanding of the consequences of environmental change. This requires an approach that

treats species as dynamic systems with plastic responses to change rather than as static entities. As such, phenotypic plasticity on an individual level and genotypic change as a population level response should be taken into account when studying the consequences of a changing world. Using a multidisciplinary approach, we report on recent advances in our understanding, identify some major gaps in our current knowledge, and point towards rewarding approaches to enhance our understanding of how environmental change alters trophic interactions and ecosystems.

INTRODUCTION

antibiotic resistance to drugs. Human-induced impacts on ecosystems comprise many facets such as the change of climatic conditions, the voluntary and involuntary introduction of non-native species and genetically modified organisms, alterations in land use, and the destruction, degradation, isolation, and fragmentation of habitats. Here we focus on evolutionary processes that result from invasive species as well as from habitat loss and fragmentation and limit our scope to these few facets of human-induced environmental change because

Key words: evolutionary processes; phenotypic plasticity; genotypic change; trophic interactions; invasive species; habitat fragmentation.

Humans can be seen as the worldÕs greatest evolutionary force (Palumbi 2001). Our actions can dramatically accelerate evolutionary processes in other species, as exemplified by the well-known induction of insect resistance to pesticides and

Received 24 September 2003; accepted 30 April 2004; published online 9 September 2005. *Corresponding author; e-mail: [email protected]

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they are considered to be the most immediate and major current threats to biodiversity (Mooney and Cleland 2001; Vitousek and others 1997). The aim of this paper is to briefly outline the state of knowledge of human-induced changes in landscapes for evolutionary processes in species communities and trophic interactions, to identify knowledge gaps and to suggest future research approaches. There is a great scientific and societal need to understand how modification and fragmentation of space affect ecological and evolutionary processes (Hunter 2002). Understanding how human-induced environmental change affects trophic interactions requires an – often lacking – integrated ecological and evolutionary approach. In the case of habitat fragmentation, for example, ecologists have traditionally investigated interactions among organisms within local assemblages, whereas evolutionary biologists studied the effects of genetic erosion of populations. Most such evolutionary studies focussed on a single trophic level. Yet, interactions occur at multiple scales, and species loss through environmental change is unevenly distributed among trophic levels, and thus modifies or disrupts the quality and quantity of interactions between organisms at different trophic levels, ultimately altering the structure of species communities. Furthermore, evolutionary processes have long been seen as long-term processes, whereas rapid, present-day evolutionary change has only recently been acknowledged as a major response to human-induced environmental change (Stockwelle and others 2003).

INVASIVE SPECIES PROCESSES

AND

EVOLUTIONARY

The rapid evolutionary responses in (and caused by) introduced species are excellent examples for the prevalence of fast contemporaneous evolution. The competitive superiority of exotic species in novel habitats make invasive species one of the largest threats to biodiversity (Mooney and Cleland 2001).

Ecological vs. Evolutionary Forces Traditionally, invasiveness has been regarded as an ecological phenomenon. Introduced species are released from fitness constraints in their native habitat, allowing some species to become invasive in their novel environment. Release from native natural enemies in their invaded ranges is the most widely accepted ecological force and there is indeed substantial support for this both in plants (Mitchell

and Power 2003, but see Agrawal and Kotanen 2003) and animals (Torchin and others 2003). However, recent studies have accumulated evidence that in many cases invasiveness evolves after colonization, rather than being an intrinsic property of the introduced species, and that evolution is integral to the study of invasion biology (Ellstrand and Schierenbeck 2000; Lee 2002; Muller-Scha¨ rer and Steinger 2004; Sakai and others 2001). Invasions often involve rapid evolutionary change (Reznick and Ghalambor 2001), facilitated by high additive genetic variance, epistasis, hybridization, the action of only a small number of genes and genome rearrangements (Lee 2002). In many species, like cordgrass, a hybridization event in the novel habitat has preceded invasion (Ellstrand and Schierenbeck 2000), which may explain the substantial time lag that is often observed between first introduction and actual invasion. In other species, selection or drift has facilitated invasion success, for example, by affecting traits involved in dispersal and species interactions. For instance, studies on populations of the butterfly Pararge aegeria suggest that morphological changes increasing flight ability accompanied colonization (Hill and others 1999). Velvetleaf (Abutilon theophrasti), introduced from Asia to North America, evolved a plastic growth response to light quality in areas with soybean, giving it a competitive advantage (Weinig 2000). In general, enemy release in the novel habitat allows introduced species to shift investment from traits involved in defence to traits that enhance growth and competitive ability, important attributes for invasiveness (the Evolution of Increased Competitive Ability, EICA hypothesis, (Blossey and No¨ tzold 1995)). At least in some ranges of their invaded habitat plants indeed appear to be larger than in their native range (Jakobs and others 2004; Thebaud and Simberloff 2001; Vila and others 2003). In several cases such differences are based on evolved genetic changes, as in the introduced Sapium tree (Siemann and Rogers 2001) and in California poppies (Leger and Rice 2003, but see Willis and others 2000) for non-genetic responses. Future studies should determine whether it is indeed selection for reduced allocation to defense that has driven such increases in size or competitive ability. So far, the expected accompanying reduction in resistance or tolerance has been documented for cordgrass populations introduced in ranges with low incidence of a specialist grasshopper common to the native environment (Daehler and Strong 1997; Garcia-Rossi and others 2003) and for introduced Sapium populations in North America (Siemann and Rogers 2003) but not for

Consequences of Invasion and Fragmentation introduced Solidago (Van Kleunen and Schmid 2003).

Genetic Change vs. Phenotypic Plasticity Evolution of interspecific interactions (trophic links, defense, loss of defense in absence of interaction, change in outcome of interaction) can occur very rapidly (Thompson 1998). However, controversy exists as to whether changes in defense traits and interspecific interactions more strongly depend on the ability of invaders and natives to genetically respond to drift and natural selection (Lee 2002), or to plasticity and broad physiological tolerance (Agrawal 2001). Thompson (1998) gives some clear examples of genetic responses, such as the rapid changes in prey defenses after contact with introduced predators. Similarly, herbivores like the soapberry bug Jadera haemotoloma rapidly genetically adapt to introduced hosts, evolving shorter stylets on the introduced small-fruited Asian goldenrain tree than on the native balloon vine (Carroll and others 2001), that is, introduced species induce evolutionary changes in native species. In some cases genetic drift rather than selection promoted successful invasions. A population bottleneck in introduced populations of the invasive Argentine ant Linepithema humile resulted in reduced genetic diversity, associated with reduced intraspecific aggression; leading to a more widespread occurrence (Tsutsui and others 2000), but such cases are probably rare. On the other hand, adaptive phenotypic responses in species interactions are widespread (Young and others 2003) and some alleged cases of evolutionary changes following introductions and altered interspecific interactions are in fact due to phenotypic plasticity (Agrawal 2001). For instance, the rapid increase in shell thickness of the intertidal snail Littorina obtusata in the Gulf of Maine following invasion by the molluscivorous crab Carcinus maenas from Europe could be explained by plastic responses to the predator (Trussell and Smith 2000) rather than to genetic change following selection by the invading predator (Seeley 1986). Perhaps, many species will respond through both, phenotypic plasticity and genetic changes, like the saltcedar does in the invasion of the cold climates of North America (Sexton and others 2002). Responses to novel environments by phenotypic plasticity offer the advantage of a rapid and often reversible response. If reciprocal phenotypic changes occur as a result of adaptive plasticity in both counterparts of a species interaction, then coevo-

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lution will result in increased phenotypic plasticity rather than in fixed adaptations and evolutionary arms races (Agrawal 2001). A future challenge is to find out whether some successful invaders are indeed more phenotypically plastic than unsuccessful invaders. Not only genetic changes but also the generally more rapid phenotypic changes in species interactions may have clear cascading effects through the food web, resulting in community level consequences (Agrawal 2001).

Ecosystem Effects Both the ecological and evolutionary changes that accompany biological invasions can dramatically alter ecosystems and their functioning by altering resource availability, food web structure, community interactions, chemical composition and the physical structure of the ecosystem itself (Crooks 2002; Vitousek 1990). Classical examples are the introduction of the fire weed Myrica faya into native ecosystems in Hawaii which until then lacked plant species with nitrogen fixing symbionts. The invading shrub enabled vegetation growth on nitrogen-poor soil and had a direct impact on the biogeochemical cycling within the invaded ecosystem (Vitousek and Walker 1989). Introduction of the brown tree snake Boiga irregularis as a top predator led to rapid extinction of the forest avifauna on the island of Guam (Savidge 1987). The cordgrass Spartina anglica, a hybrid between native S. maritima and introduced S. alterni flora, invaded intertidel flats of the British Isles, replacing more diverse native plant communities and altering succession (Gray and others 1991). The Asian clam Potamocorbula amurensis, which invaded the San Francisco Bay in 1986, dramatically and persistently lowered chlorophyll biomass and reduced primary production 5-fold (Alpine and Cloern 1992). Terrestrial flatworms from New Zealand have successfully invaded Northern Ireland and substantially reduced the populations of their earthworm prey (Moore and others 1998), with potential dramatic negative effects on the ecosystem functions performed by the earthworms. On the other hand, earthworms themselves can destroy ecosystems when invasive: The Amazonian species Pontoscolex porethrurus deposited so much highly-compacted cast material on the soil surface that it became impermeable for both plants and water (Chauvel and others 1999). Plants producing toxic substances such as the bracken fern (Pteridium aquilinum) often become invasive due to their high competitive strength and low sensitivity to herbivory and can change the community structure of

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whole ecosystems by displacing the originally dominant species, for example, Calluna vulgaris (Anderson and Hetherington 1999). Their studies also revealed that in transition zones with both species occurring simultaneously, litter decomposition rates might be considerably enhanced. In the long term, this would result in a number of soilrelated alterations such as a reduced litter layer and humus content, with concomitant effects on the inhabiting organism community and the manifold functions performed by them. Besides invasive species altering ecosystems, their invasive potential is also affected by ecosystem properties: Plankton food webs were more sensitive to invasion by the exotic Daphnia lumholtzi when they had a low nutrient status, accompanied by low zooplankton biomass and high zooplankton species diversity (Lennon and others 2003). Where invaders cause, for example, increased herbivory of plants that produce volatile compounds as a defence strategy, this will result in higher energy demand of the affected plants (Hoballah and others 2004). The authors, however, conclude from earlier studies that these costs will be counterbalanced by the beneficial effects of volatiles as long as natural enemies of the herbivores are present in the respective ecosystem. Besides nutritional aspects, the extent to which the genetic diversity of an invaded community affects invasibility has been largely unexplored. An experiment with varied numbers of Arabidopsis thaliana genotypes being invaded by Arabidopsis suecica revealed no effect of genetic diversity on invasibility, instead A. thaliana density had strong effects on size and reproduction of A. suecica (Weltzin and others 2003).

FRAGMENTED LANDSCAPES AND EVOLUTIONARY PROCESSES Ecological vs. Evolutionary Processes Habitat destruction and the fragmentation of the remaining habitat area can lead to dramatic losses of biodiversity (Saunders and others 1991). Because the spatial scale of population processes and species-area relationships depend on trophic rank (Holt and others 1999), decreased size and connectivity between habitat fragments often disrupts trophic interactions between plants and their pollinators, competitors and natural enemies (Tscharntke and Brandl 2004). However, although a number of studies has provided insight into how species interactions are affected by the differential effects of fragmentation on the abundance and dynamics of species, evolutionary changes in spe-

cies interactions following habitat fragmentation are still not widely appreciated (Ouborg and Biere 2003). Habitat fragmentation often results in increased rates of inbreeding and genetic drift (Ellstrand and Elam 1993). The increased levels of inbreeding can affect the persistence of individual species directly (Keller and Waller 2002), but can also affect their biotic interactions through alteration of their resistance or tolerance to natural enemies (Ouborg and Biere 2003). For instance, inbred monkey flowers (Mimulus guttatus) are more susceptible to spittlebugs than outcrossed plants, suffering a three-fold higher reduction in biomass and flower production (Carr and Eubanks 2002). Although the precise effects of inbreeding on resistance often vary within and among populations, depending on parental genotype, inbreeding history and covariance structure of traits underlying resistance (Strauss and Karban 1994), such studies indicate that fragmentation can alter species interactions through effects on the genetic population structure of the species involved in the interaction. In addition, habitat fragmentation may lead to evolutionary changes in species interactions due to responses of species to changes in the encounter rate with their mutualists, competitors, natural enemies, or prey (Bolger and others 2000). Small size and isolation of plant populations often reduces the probability of colonization and spread of infectious diseases and herbivores (Simberloff 1988). Hess (1994) even showed that due to increased likelihood of disease transmission, increased connectivity between habitat fragments could enhance the probability of metapopulation extinction and reduce patch occupancy, although often the benefits of corridors are likely to outweigh risks of disease transmission (McCallum and Dobson 2002). In turn, the changes in encounter rates may affect the life history evolution of the organisms involved. For instance, changes in encounter rate with herbivores and pathogens will affect allele frequencies at loci involved in pathogen and herbivore resistance (Hochberg and Mo¨ ller 2001). Likewise, fragmentation may reduce the encounter rate with beneficial organisms such as pollinators (for example, Goverde and others 2002) and vertebrate seed dispersers (for example, da Silva and Tabarelli 2000). This may lead to extinction cascades of plants and their associated herbivores and upper trophic level predators and parasitoids. Yet, the low abundance of pollinators in habitat fragments can also result in selection for reproductive assurance in plants, for example, through partial inbreeding (Inoue 1993).

Consequences of Invasion and Fragmentation In conclusion, predictions of the effects of habitat fragmentation that are based on spatial ecology (decreased encounter rate with natural enemies and mutualists) may substantially and qualitatively differ from predictions based on short term population genetic consequences, such as increased susceptibility upon a concomitant increase in inbreeding, and longer term evolutionary consequences (altered selection for traits related to resistance and dispersal). Hence, an integrated approach is necessary to predict changes in trophic interactions upon habitat fragmentation.

Effects of Spatial Behaviour Any effect of habitat fragmentation depends on the – often species specific – spatial scale experienced by an organism (Thies and others 2003) and on a number of other factors. High trophic level, large body size, high habitat specificity and low dispersal capacity are species traits that have often been associated with sensitivity to fragmentation (Kruess and Tscharntke 2000; Tscharntke and Brandl 2004). This suggests at first sight that plants in fragmented landscapes should be released from herbivore pressure. Yet, because predators or parasitoids of herbivores often suffer even more from fragmentation due to low dispersal abilities, herbivores might be released from their enemies, and thus the attack rate on plants might become even higher. In contrast, in a large data set comprising fragments varying in size and age, Bolger and others (2000) found spiders and carabid beetles to increase in fragmented habitats, especially in small fragments – which would result in increased predation rates. In this study, invasive ants (Linepithema humile) and other non-native arthropods also increased in abundance in such areas, and it is reasonable that the observed increase in spiders and predators was not a direct effect of fragmentation but rather resulted from increased prey availability. In the same vein, Bruhl and others (2003) report an increase of tramp species with decreasing fragment size. Predators can also affect the spatial behaviour of their prey: migration patterns of sandpipers are influenced by falcon predation, with possible consequences on the evolution of migration routes (Lank and others 2003). At high risk of predation, dusky warblers (Phylloscopus fuscatus) preferred safe nest sites over those offering other advantages (Forstmeier and Weiss 2004). Van Nouhuys and Hanski (2002a) provide an interesting example of two insect parasitoids that attack the same host, but differ significantly in their dispersal abilities. Here,

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habitat isolation disrupts the competition between both parasitoid species and the tritrophic interaction between plant, herbivore, and parasitoid. These examples illustrate the importance of knowing the resource requirements, behaviour and life-history of species, if we are to predict the consequences of habitat fragmentation and, possibly, also of other human-induced environmental changes. It is noteworthy that the cause of fragmentation also has to be considered when deriving consequences for evolutionary processes: for example, the phylogeography of Amazonian frogs and small mammals could not be assigned to a riverine barrier, but only to the thrust-and-fold dynamics of the Andes (Gascon and others 2000). Overall we know very little of spatial behaviour in general. Important questions are: how does scale, for example patch size and patch connectivity, affect the population dynamics of trophically linked species, particularly in light of highly variable life history strategies? How do fine-scale movement patterns translate to broad-scale species distribution patterns? And what is the impact of scale on evolutionary dynamics? Singer and Thomas (1996), for example, showed that scale governed the relative roles of behavioural processes (such as host preference) and ecological processes (such as resource-specific population growth) in determining the distribution of the butterfly Euphydryas editha across resources. As yet, we are largely ignorant of the contemporaneous evolution taking place in multitrophic systems in fragmented landscapes. Adaptation to environmental change can occur through behavioural modification (phenotypic plasticity) or life-history evolution. For invertebrates (and probably also birds) patch choice, patch residence times, diet choice (host selection) and clutch size are sensitive to changes in travel time between patches (Stephens and Krebs 1986). Therefore, changes in travel time and patch or host selection can affect the spatial heterogeneity in attack rate by enemies, which is a significant factor in the persistence of host-parasitoid and predator-prey interactions (Bernstein and others 1991). Hence, there is a clear need to study how organisms perceive their environment and behaviourally deal with it (Vet 2001). Without this insight we can only guess at the evolutionary and ecological consequences of change in landscape structure for species communities.

Effects of Habitat Type Both ecological and evolutionary processes in response to environmental change might greatly

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differ depending on the type of habitat. Poor habitat quality hampers population growth, extending to associated density effects (Rotem and Agrawal 2003; Ruohoma¨ ki and others 2003). Birds are mobile enough not to be directly influenced by fragmentation, but their food is affected by both habitat quality and fragmentation. Low food availability will result in low accumulation of fat stores, which have been shown to affect the departure of barnacle geese (Branta leucopis) for migration (Prop and others 2003). This is a clear example of how spatial behaviour can be affected by habitat quality. Habitat properties can also affect sexual selection: the size and availability of appropriate nest sites determined habitat preference, intrasexual competition and clutch size in sand gobies (Pomatoschistus minutus) (Lehtonen and Lindstro¨ m 2004). Simulation models revealed strong effects of patch heterogeneity and movement decisions on population development (Russel and others 2003). Also patch size and isolation are an essential prerequisite for understanding insect distribution patterns at multiple spatial scales (Krawchuk and Taylor 2003). The significance of habitat configuration is not uniform for all types of organisms, though. In a field experiment, soil organisms were strongly influenced by habitat quality but only weakly by fragmentation: most such organisms are very small and adapted to small-scale, highly patchy environments, thus not sensitive to fragmentation (Rantalainen 2004).

plant nitrogen acquisition (Wardle and others 1997). Habitats with an evolutionary history of low disturbance frequency have favoured the evolution of specialists that should be especially prone to suffer from habitat fragmentation and isolation. For example, anthropogenic disturbance has led to grave reductions in guilds of fish species that are specialized to specific riverine conditions, with rheophilic fish species becoming especially rare due to the degradation and fragmentation of their reproductive habitats (Aarts and others 2004). Extinctions on lower trophic levels here clearly lead to extinction cascades on dependent higher trophic levels. Habitats with high disturbance frequencies in their evolutionary history, in turn, will inhabit generalists at higher frequencies that can switch to alternate resources if one of their resources is driven to extinction. In general, strong and rapid environmental change will support generalists at the expense of specialists (for example, see Warren and others 2001). In the case of plant herbivore interactions, the loss of specialist herbivores and concomitant increase in prevalence of generalists should lead to the evolutionary process of a shift in plant defence that is similar to that observed in invasive plants (the EICA hypothesis, see above). This is likely to impoverish the potential number of coexisting species, further eroding biodiversity.

Effects of Disturbance Frequency and Extent

NOVEL APPROACHES TO COMBINE ECOLOGICAL AND EVOLUTIONARY APPROACHES

Disturbance is a key issue in habitat properties that determines occurrence and performance of both native and invasive species in ecosystems. For example, the bracken fern Pteridium aquilinum especially invades and persists in areas with high frequency of fire, not only due to its ability to sprout quickly after a fire and preference for alkaline soils, but also by producing highly flammable litter. Johst and Drechsler (2003) have shown that disturbance is particularly negative for species survival when the single disturbance events are spatially correlated. Habitat fragmentation can alter disturbance frequency (White and Jentsch 2001) and thus potentially has strong ecosystem effects. For example, island area has important effects on the disturbance frequency of forest fires, inducing area related differences in species composition, standing biomass, plant litter decomposition, nitrogen mineralization, terrestrial carbon partitioning, humus accumulation, and

In a recent special feature of Ecology, a strong plea was made for Ôcommunity geneticsÕ as a synthesis between community ecology and population genetics, to yield new insights into the interplay between evolutionary and ecological processes (Agrawal 2003). On the one hand this will help us to understand the evolution of organisms, which occurs in a community context. On the other hand it can elucidate the role of genetic variation within a species at higher levels of biological organization. Habitat fragmentation is one of the prime examples where such an approach can be useful to understand the community-wide impact of an anthropogenic disturbance, as fragmentation leads to both genetic and demographic changes that affect the abundance of species as well as the trophic interactions they are involved in. In the same sense, Ôlandscape geneticsÕ, that is, the combination of landscape ecology and population genetics, has emerged as a promising approach to facilitate our

Consequences of Invasion and Fragmentation understanding of how geographical and environmental features structure genetic variation at the population and individual level (Manel and others 2003). Although both approaches involve the population genetic level, we suggest the contemporaneous evolution of life-histories in changing environments as an additional field of research that should yield important new insights. This rapid form of evolution has been documented in invasive plants, and a promising approach here is to combine landscape ecology with evolutionary processes (With 2002) to predict the invasive potential of non-native plants. In contrast to invasive plants, the role of contemporaneous evolution in multitrophic interactions is largely unknown (but see Yoshida and others 2003). In predators and parasitoids that hunt for prey and hosts in fragmented landscapes, limited resources may either be invested in fecundity or in somatic maintenance (Ellers 1997). Increasing fragmentation might result in disruptive selection for individuals with either low fecundity and increased dispersal ability or high fecundity and low dispersal ability. How strongly such selection operates will depend to some extent on the matrix surrounding the habitat fragments. Although classical metapopulation ecology treats the matrix as uninhabitable (Wiens 1997), recent investigations have shown that some effects of habitat fragmentation on species persistence depend greatly on the complexity of the surrounding landscape (Cook and others 2004; Murphy and Lovett-Doust 2004; Tscharntke and others 2002) and on the way organisms behaviorally respond to this complexity. Depending on the specific behavior of the species, landscape structure may help or impede connection of isolated patches. For insects, the role of habitat corridors for connectivity is in great need of further study (Hunter 2002). Recent studies found little evidence for corridors to foster soil decomposer arthropods, whereas enchytraeid worms profited from their presence (Rantalainen 2004).

CONCLUSIONS Our understanding of how environmental changes, especially the introduction of non-native species, as well as the fragmentation and isolation of habitats affect species communities has greatly improved over the past decade. Today we have a reasonably good understanding of rapid evolutionary processes taking place in invasive species that lead to increased competitive ability. Less well understood are defence-specific trait trajectories, the role of the

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evolutionary history and the degree to which the novel habitat is disturbed. Reciprocal transplant experiments can be used in combination with phenotypic or genotypic selection analysis to examine the direction and strength of selection in native and novel habitats (Mu¨ ller-Scha¨ rer and Steinger 2004). Equally improved is our knowledge with respect to what kinds of organisms are especially threatened by habitat fragmentation (for example, Tscharntke and Brandl 2004; Tscharntke and others 2002), and ecological theory, notably metapopulation theory, has recently advanced to study metapopulation effects on communities rather than single species (Holt 1997) and the interactions among those communities (Van Nouhuys and Hanski 2002b). However, these approaches have been purely ecological ones. Evolutionary biologists, in contrast, have largely focussed their interest on genetic drift and genetic erosion in fragmented populations, and until very recently, there was no unified approach, combining evolutionary and ecological aspects. The novel fields of community genetics (Agrawal 2003) and landscape genetics (Manel and others 2003) will certainly yield new insights and amalgamate ecological and evolutionary approaches. Studying entire communities and their functioning in ecosystems is almost impossible – therefore it will be necessary to first focus on (a selection of) keystone species within each ecosystem, and on traits that make them particularly sensitive to fragmentation and/or invasion. We propose a set of criteria (Table 1) according to which each species in question can be scored and that eventually allows for selecting those species that are most likely to be affected by fragmentation/invasion, with concomitant effects on the whole ecosystem. As fragmentation or invasions in most cases will not result in species extinction but rather in more or less dramatic changes in the dominant structure, it is important to keep in mind the effects of population density: interactions in most species are non-linearly density dependent. Changes in population density can affect production, population growth and reproduction (Rotem and Agrawal 2003; Ruohoma¨ ki and others 2003) or cause physiological or behavioral changes (for example, cannibalism, Wagner and Wise 1996) that are associated with energy demand, that is, the performance of the whole system. One of the most prominent examples of density effects is the mass migration of locusts caused by crowding and accompanied by morphological changes (Schowalter 2000).

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Table 1. Criteria for Selecting a Set of Species when Studying Potential Effects of Habitat Fragmentation or Species Invasion on (Evolutionary Processes in) Ecosystems Criterion

1

2

3

Keystone species Abundance Frequency Body size Dispersal distance Habitat specifity/niche width Trophic position Feeding strategy/nutrient requirements Reproductive strategy Competitive strength Generation time Sensitivity to disturbance

definitely high high large low high/narrow predator, parasite specialized, high sexual, specializeda low long high

probably medium medium or variable medium medium medium consumer, detritivore polyphagous, intermediate sexual medium intermediate intermediate

probably not low low small high low/wide producer, saprophyte omnivorous, low parthenogenetic high short low or favored by disturbance

Species with most hits in column 1 should be preferably studied because they are most likely to be affected and/or effects on them will result in trophic cascades and alterations of ecosystem performance. a E.g., specific requirements for mating behavior or nesting

An important area in which ecology and evolution can meet is in assessing the role of movement and across-spatial scale processes on the ecology and evolution of species and community structure. Fragmentation may select for strong dispersal ability in species that are able to respond rapidly, possibly dampening the impact of fragmentation on the dynamics of species (Dubbert and others 1998). Changes in dispersal ability may induce adaptive changes in foraging behaviour and phenotypic plasticity may turn out to be a pervasive adaptive trait to respond to habitat modification. When species evolve greater dispersal ability, we may simultaneously expect evolutionary changes in the life history of species (Hughes and others 2003), due to the basic life history trade-off between somatic maintenance and fecundity. If fragmented habitats select for species with high dispersal rates, this should result in greater uniformity of ecosystems, resulting in lower b-diversity and eventually a general loss of biodiversity. This area of linking spatial ecology to evolution will certainly offer new and important insights for an understanding of how changes in landscape structure affect multitrophic interactions in species communities. Combining ecological and evolutionary approaches as described above should allow us to make more accurate predictions of how species will respond to a large anthropogenic impact such as habitat fragmentation and hence improve our management and conservation of biodiversity. Some of the areas that represent particular gaps in our current knowledge include: (1) To what

extent can increased dispersal, evolutionary change and plasticity ameliorate effects of landscape change and species introductions? (2) Are organisms and evolutionary processes in particular types of landscape (for example, young vs. mature ecosystems) more sensitive to landscape change and species introductions than those in other landscapes? (3) What are the ecosystem-level consequences of evolutionary shifts in species interactions following landscape change and species introductions? (4) What is the significance of density-dependent effects in ecosystems affected by fragmentation or invasions? (5) Is there a pervasive role for plasticity in altering species interactions following landscape change and species introductions? (6) Is phenotypic plasticity a key trait in invasive species? (7) Which keystone species (in specific ecosystems) are particularly at risk from fragmentation or invasions? We hope this paper will stimulate research that will help close some of the gaps in our understanding of evolutionary processes in trophic interactions and ecosystems that result from environmental changes such as the invasion of non-native species and the fragmentation of habitats.

ACKNOWLEDGEMENTS This paper is based upon a discussion group at a European Science Foundation exploratory workshop at Texel, NL, that was organized by Wim van der Putten, Peter de Ruiter, and Martin Wassen. We are grateful to Wim van der Putten, Jeffrey Harvey and two anonymous referees for valuable comments on earlier versions of this manuscript.

Consequences of Invasion and Fragmentation

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