In this section, we review the rather well-developed theory regarding the effects of habitat fragmentation on population viability, and we evaluate whether the available empirical evidence for forest plant species fits these theoretical predictions.
1. Population genetic consequences of small population size and isolation
Increased spatial isolation and decreased population size may lead to the erosion of genetic variation and increased genetic differentiation among populations through genetic drift, increased inbreeding and reduced gene flow between populations (Schaal & Leverich, 1996; Young et al., 1996). Genetic drift is the random change in allele frequencies because alleles transmitted from the parental population to the offspring are a random sample of all available alleles. Genetic variation in small populations is likely to be reduced and allele frequencies become highly unpredictable due to genetic drift. Because the effect of genetic drift on genetic diversity largely depends on the number of generations during which the population remains small (Young et al., 1996), it can be expected that the loss of variability through drift is relatively small in forest plant species with long generation times, prolonged clonal growth, low seed production and limited seedling recruitment. Unfortunately, there are few population genetic studies dealing with herbaceous forest plant species that allow testing this prediction. Reduced allelic diversity was found in small populations of Viola pubescens and Primula elatior, both of which show repeated seedling recruitment (Van Rossum et al., 2002; Culley & Grubb, 2003). Jacquemyn et al. (2004), however, could not confirm this trend for P. elatior. Neither the strongly clonal and rarely recruiting perennial Maianthemum bifolium, nor the perennial Microseris lanceolata, showed a relation between population size and population genetic diversity (O. Honnay, University of Leuven, unpublished results; Prober et al., 1997, respectively). Young et al. (1996) suggested that species with long generation times and low seedling recruitment lose genetic diversity mainly through founder effects and population bottlenecks, in other words sudden reductions in population size. This is also what Tomimatsu & Ohara (2003) invoked to explain the low genetic diversity observed in small and isolated populations of the perennial Trillium camschatcense in Japan. Reduced genetic diversity may decrease the potential of a species to adapt to (slow) environmental change. In the relatively short term, reduced genetic diversity in a population may contribute to increased inbreeding and to lowered levels of heterozygosity. In small populations, the probability of bi-parental inbreeding or inbreeding through increased self-fertilization increases. Increased homozygosity is expected to directly affect fitness through inbreeding depression (Ellstrand & Elam, 1993). Studies that related forest fragmentation with decreased heterozygosity and reduced fitness are rare, however. Neither Tomimatsu & Ohara (2003) nor Van Rossum et al. (2002) found an effect of population size on heterozygosity for the forest plants T. camschatcense and P. elatior, respectively. On the other hand, Culley and Grubb (2003) found a heterozygosity deficit in small (< 300 individuals) populations of V. pubescens. This deficit was attributed to increased inbreeding, mainly by increased self-pollination in small and isolated populations due to a decline in the abundance of solitary bees, the main pollinator of the species, in small and isolated forests. This example shows the interrelatedness of fragmentation-related genetic effects and effects of changes in the pollinator community. The latter will be dealt with in Section IV.3.
Finally, increased fragmentation will increase genetic differentiation between populations (Young et al., 1996). In the absence of gene flow (through seeds and/or pollen) between fragmented populations, and given a stable population structure, random genetic drift will increase genetic differentiation among populations (Schaal & Leverich, 1996). Because most herbaceous forest plants are insect pollinated and because most pollinating insects do not travel large distances (Wilcock & Neiland, 2002), especially when they have to cross a hostile landscape matrix, the proportion of pollen flow in the total gene flow between isolated populations is probably rather low. Consequently, long-distance gene flow between fragments occurs mainly by accidental seed transport through endo- or exo-zoochory (e.g. Vellend, 2003). A very low level of gene flow (c. 1 recruiting seed per generation) is sufficient to reduce genetic differentiation due to genetic drift (Wright, 1931). High levels of genetic differentiation (30–40%) were found for Anemona nemorosa, Viola riviniana and V. pubescens (Stehlik & Holderegger, 2000; Auge et al., 2001; Culley & Grubb, 2003, respectively). Although in some cases the studied species exhibited long lifetimes (genets of more than 200 yr old were discovered in A. nemorosa), all these species are characterized by extensive seed set and repeated seedling recruitment, which are enabling processes of genetic drift. Genetic differentiation was found to be much lower (c. 10%) for the strongly clonal M. bifolium, characterized by very low seedling recruitment and extreme seed dispersal limitation, and in small populations of T. camschatcense (Tomimatsu & Ohara, 2003). The perennial and abundantly recruiting P. elatior, exhibiting very good seed dispersal between fragmented forests, showed low population differentiation (c. 5%) (Van Rossum et al., 2002; Jacquemyn et al., 2004).
The available studies show how dispersal between fragments on the one hand and long generation times and prolonged clonal growth on the other hand may interact in determining population genetic differentiation (Table 1). The highest reported percentages of population differentiation (30–40%) for forest plants corroborate the average percentages for long-lived perennials and species of late successional habitats reported by Nybom and Bartish (2000). These percentages are still well below the reported range of population differentiation for annuals and short-lived perennials (40–70%).
Table 1. Two counteracting processes determine the degree of genetic erosion and population differentiation of forest plant populations in fragmented landscapes
| || ||Sensitivity to small population size|
|LOW Long generation time (limited seedling recruitment)||HIGH Short generation time (repeated seedling recruitment)|
|Sensitivity to isolation||HIGH||Low differentiation||High differentiation|
|Low seed dispersal capacity||(Maianthemum bifolium,Trillium camschatcense)||(Anemone nemorosa, Violariviniana, Viola pubescens)|
|High seed dispersal capacity|| ||(Primula elatior)|
As a general conclusion, we can say that for at least some species, even small populations may not have lost alleles, and that they may still contain high levels of genetic variability. It appears that knowledge of the degree of seedling recruitment, the extent of clonal reproduction and the seed dispersal capacity may give a first indication of the susceptibility of the populations to genetic erosion and genetic differentiation (Table 1). It is clear that this trend may be complicated by the degree of self-compatibility of the species and, although we excluded genetic studies covering a geographical gradient (e.g. Griffin & Barrett, 2004), also by the degree of spatial isolation of the populations.
2. Decreased mate availability in self-incompatible forest plant species
Self-incompatibility is the failure of a fertile hermaphroditic seed plant to produce zygotes after self-pollination (De Nettancourt, 1977). Almost one third of all forest plant species show a self-incompatible breeding system and almost all of them show prolonged clonal growth. The latter can strongly affect the number of genotypes in a population (for an extensive treatment, see Honnay & Bossuyt, 2005). Certainly under unfavourable environmental conditions such as low light availability through canopy closure, forest plant species may alter their growth form and exhibit prolonged clonal growth (e.g. Kudoh et al., 1999; Lezberg et al., 2001). The direct consequence of prolonged clonal reproduction and suppression of sexual reproduction is that locally less-adapted clones become outcompeted by expanding ramets of more adapted genotypes, ultimately leading to populations with very few genotypes (Hartnett & Bazzaz, 1985; Eriksson, 1989). When the environmental pressure is diminished, for example by the opening of the canopy through forest management, sexual reproduction will be hampered or even completely impossible (i.e. sexual extinction, as in Eckert, 2000) due to very low genotypic diversity and lack of compatible pollen. Such a loss of genotypes has been described for the North American prairie plant Asclepias meadii (Schaal & Leverich, 1996). A similar process has resulted in monoclonal patches of Maianthemum canadensis in North American forests (Worthen & Stiles, 1988) and lowered genotypic diversity of the understorey species Uvularia perfoliata (Kudoh et al., 1999). Thus, although prolonged clonal growth may offer an escape route from genetic drift, the consequences in the long term may be complete sexual extinction of the population. Although the described process is initially more related to forest management than to fragmentation, it can be expected that genotype loss will occur faster in isolated populations because of the lack of addition of new genotypes through seed inflow (Schaal & Leverich, 1996).
A second mechanism leading to sexual extinction in prolonged clonally propagating species is summarized by the Somatic mutation theory of clonality (Klekowski, 1988, 1997), which states that sexual reproductive success is inversely proportional to longevity. The older and larger a clone becomes, and the longer the periods between sexual reproduction between clones, the more likely that mutations will be accumulated that decrease the probability of successful sexual reproduction in the population (Lamont & Wiens, 2003). Mutations inducing infertility will contribute to an acceleration of the above described process towards monoclonal patches because the number of available fertile genotypes for pollination further decreases. The precise role of mutations in sexual sterility, however, remains an open question (Eckert, 1999), but long-lived forest plants can be expected to be very susceptible. All of this suggests that forest management and the spatial isolation of populations may interact with and affect the genetic structure of forest plant populations. More work on the relation between forest management, clonal growth, isolation and genotypic diversity is necessary.
3. Changed interactions with pollinators in small and isolated populations
Most forest plant species show clonal propagation to some extent (see Section IV.2), but almost all of them also reproduce sexually. Sexual reproduction in forest plant species is important because it helps maintain genetic diversity within populations and because only seeds (or fruits containing seeds) are capable of long-distance dispersal, which may be essential in continuously changing landscapes (see Section VI.2). Like most other plant species, forest plant species also rely on animals (especially insects) for effective pollination and sexual reproduction. Fragmentation may negatively affect pollinator abundance and diversity in habitat fragments (Kearns et al., 1998). Fragments can become too small to sustain pollinator communities or too isolated to attract a large diversity of pollinators (Steffan-Dewenter & Tscharntke, 1999), both of which affect pollinator efficiency and therefore the reproductive success of plant species. In addition to these direct effects of fragmentation through changes in pollinator guilds, indirect effects associated with altered pollinator behaviour and flight patterns within patches have been shown to affect reproductive success of forest plant species in fragmented forests (Didham et al., 1996). Patterns of pollinator behaviour depend on the number and density of individuals in a population (Kunin, 1993, 1997). The relative importance of direct and indirect effects of fragmentation on pollination efficiency and hence on reproductive success can be expected to vary greatly among species, depending on the degree of dependence on pollinators for successful reproduction (generalist vs specialist species) (Bond, 1994; Waser et al., 1996) and on the breeding system (Aizen et al., 2002).
Although studies investigating the effects of forest fragmentation on pollinator community structure are very rare, the few available studies point to a clear decreased pollinator diversity and abundance in small forest fragments (Murcia, 1996). Aizen and Feinsinger (1994a) showed that both the abundance and species richness of native euglossine bees significantly declined in Argentinean small (< 1 ha) forest fragments compared to large (2–6 ha) fragments and continuous forest. Forest fragments were situated in an intensively used matrix of agricultural land, unsuitable for pollinating insects. Further investigation of reproductive success of 16 plant species in these forest fragments showed that 81% of the plant species exhibited a significant decline in pollination and 73% exhibited a significant decline in seed set in the small fragments compared with the continuous forest (Aizen & Feinsinger, 1994b). The median decrease in the level of seed set between different species between small fragments and continuous forests was 20%.
In addition to changes in pollinator community structure, the small population size associated with the small patch sizes makes populations rather inconspicuous for pollinators, leading to lower visitation rates and hence decreased reproductive success. Sih and Baltus (1987), for example, found strongly reduced reproductive success in small populations of Nepeta cataria at the edges of woodlands. Similar results were obtained by Jacquemyn et al. (2002) and Aguilar and Galetto (2004), who showed decreased reproductive success in small populations of the forest herbs Primula elatior and Cestrum parqui, respectively. In both cases, a decreased number of visits in small populations partly explained the lower reproductive success. In the case of P. elatior, reduced mate availability strengthened the effects of small population size. This distylous species has two style morphs: a long (pin) and a short-styled (thrum) morph (Jacquemyn et al., 2002). Successful reproduction is only possible between the two different morph types. In very large populations, it can be statistically expected that the morph occurrence is 50%. In very small populations, however, chance effects will bias morph type occurrence and hamper or even prevent successful reproduction (Fig. 2a,b; see also Endels et al., 2002). Similar effects of limited mate availability on plant reproduction were demonstrated for the boreal forest plant Linnaea borealis (Wilcock & Jennings, 1999).
Figure 2. Reduced reproductive success in small populations of Primula elatior (a) may be at least partially explained by reduced mate availability due to a bias between short- and long-styled individuals in small populations (b) (from Jacquemyn et al., 2002).
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The strongly reduced reproductive success in small populations of the self-incompatible P. elatior may indicate that specialist species or species characterized by a self-incompatible breeding system should be more susceptible to fragmentation than generalist plant species or species with a compatible breeding system. However, this does not seem to be the case. Based on a review of 25 studies and 46 plant species, Aizen et al. (2002) found no significant relationship between the breeding system (self-compatible or self-incompatible) of a species and its response to fragmentation. This does not mean that the effects of lowered mate availability on population persistence are not important. These effects may be compensated by other habitat fragmentation-induced changes on population fitness through edge effects, pollinator limitation or genetic erosion, for example (see also Ashworth et al., 2004). This is another piece of evidence showing that generalizations on the effects of habitat fragmentation are very difficult to make and that the ultimate effects on population fitness may be difficult to disentangle.
Even though these results demonstrate direct effects of forest fragmentation on reproductive success, it is not clear whether these reductions translate into reduced recruitment rates and low population growth rates. There is some evidence that a reduced seed set may lead to decreased recruitment rates and hence to smaller growth rates. For individuals of Trillium ovatum occurring along clear-cut edges, recruitment was nearly nonexistent, and expected population declines ranged from 54% to 83%. The reduced recruitment rates at the forest edge were partly a result of a decreased seed production due to lower pollinator activity (fruiting in forest edge populations was more pollen-limited than in others) and increased seed predation (Jules, 1998; Jules & Rathcke, 1999). More research on the relationship between reproductive output and recruitment rates is needed to understand fully the impact of reduced pollination success on plant population persistence.
4. Edge effects
Forest fragmentation also implies a relative increase in edge habitat. For an identical shape, a small forest fragment has a higher edge : core ratio than a large forest fragment. Hence, edge effects become much more important in highly fragmented landscapes. It has been suggested that the effects of habitat fragmentation through edge effects may be more important than area and isolation effects per se (Turner et al., 1996; Harrison & Bruna, 1999). Edge effects may affect forest plant dynamics such as regeneration and interspecies competition and, as described above, also plant–animal interactions (predation, seed dispersal and pollination) (Murcia, 1995). The changed microclimate at the forest edge, characterized by increased light penetration, increased air and soil temperature, decreased air humidity and an increased level of agro-nutrients in the soil, directly affects population dynamics of the occurring plant species.
In temperate forests, most authors report a transient zone between the landscape matrix and the unaffected core area of the forest of between 20 and 50 m, depending on the aspect of the edge (in the northern hemisphere, edge effects are much more pronounced along south- than along north-facing edges) (e.g. Palik & Murphy, 1990; Matlack, 1993; Williams-Linera et al., 1998; Honnay et al., 2002b; and many others).
Studies focusing on changes in the population dynamics of forest plant species near edges, and thus giving direct insight in the ecological mechanism behind community composition changes, are much less common. Jules (1998) found a dramatic decrease in the recruitment in T. ovatum closer than 100 m to clearcut edges. In addition to changes in pollinator behaviour (see Section III.3), the ecological mechanisms behind this demographic change were a direct consequence of a changed microclimate in the forest edge zone. Similar results were obtained for T. camschatcense where recruitment of juvenile stages through seed germination was even so strongly limited near forest edges due to microclimate changes (Tomimatsu & Ohara, 2004). These negative edge effects on plant reproduction cannot be generalized, however. In Australian mallee woodland strips, edges were characterized by decreased predation rate and increasing undamaged fruit production of the understorey species Eremophila glabra (Cunningham, 2000). Although edge effects in general appear to affect population viability negatively by hampering successful recruitment, complex interactions between abiotic and biotic edge processes hamper generalization.
Finally, edge effects may also enhance the invasion of the forest by species not normally occurring, or naturally occurring only at low densities in forest edges. Microclimate changes which are typical for edges often give extra competitive advantages to these invasive species over the naturally occurring forest plant species. Yates et al. (2004) found significantly higher abundances of the exotics Rosa multiflora and Lonicera japonica in forest edges. The effects on the naturally occurring forest plants species were not clear, however. Brothers & Spingarn (1992) and Honnay et al. (2002b) found that limited light availability in the forest core prevented most exotics and weedy species from penetrating more than a few metres into the forest edge. Generally, forests seem quite well buffered against invasion of nonforest plants due to the rapidly decreasing light availability when entering the forest fragment. Moreover, most forest fragments still possess a relatively large undisturbed core area. The remaining unaffected core area can be calculated given the shape of the forest fragment, the area, and the average penetration distances of the edge effects (Laurance & Yensen, 1991).