Possible effects of habitat fragmentation and ... - Wiley Online Library

Ecology Letters, (2002) 5: 525–530

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Possible effects of habitat fragmentation and climate change on the range of forest plant species

Olivier Honnay*, Kris Verheyen, Jan Butaye, Hans Jacquemyn, Beatrijs Bossuyt and Martin Hermy Laboratory for Forest, Nature & Landscape Research, University of Leuven, Vital Decosterstraat 102, B-3000 Leuven, Belgium *Correspondence: E-mail: olivier.honnay@ agr.kuleuven.ac.be

Abstract Global circulation models predict an increase in mean annual temperature between 2.1 and 4.6 °C by 2080 in the northern temperate zone. The associated changes in the ratio of extinctions and colonizations at the boundaries of species ranges are expected to result in northward range shifts for a lot of species. However, net species colonization at northern boundary ranges, necessary for a northward shift and for range conservation, may be hampered because of habitat fragmentation. We report the results of two forest plant colonization studies in two fragmented landscapes in central Belgium. Almost all forest plant species (85%) had an extremely low success of colonizing spatially segregated new suitable forest habitats after c. 40 years. In a landscape with higher forest connectivity, colonization success was higher but still insufficient to ensure large-scale colonization. Under the hypothesis of net extinction at southern range boundaries, forest plant species dispersal limitation will prevent net colonization at northern range boundaries required for range conservation. Keywords Dispersal limitation, dispersal mode, forest fragmentation, global change, landscape connectivity, plant migration, range contraction. Ecology Letters (2002) 5: 525–530

INTRODUCTION

Global circulation models predict an increase in mean annual temperature between 2.1 and 4.6 °C by 2080 in the northern temperate zone (IPCC 2001). For ecologists the challenge is to predict the effects of climate warming on species and communities. It is expected that globalwarming-induced changes in the ratio of extinctions and colonizations at the northern and southern boundaries of species ranges will result in northward range shifts for many species (e.g. Saetersdal et al. 1998; Parmesan et al. 1999; Thomas & Lennon 1999). At the southern range boundary, the climate-induced modifications in the species physiology, phenology and ecology inevitably cause net extinction (e.g. by the invasion of mobile competitive species from the south which change local competitive equilibria). At the same time, at northern range boundaries, increasing temperatures will allow net colonization by enabling species to colonize new suitable habitats which are beyond their original range boundary (Cramer et al. 1999; Hughes 2000). However, net species colonization at northern boundary ranges, necessary for a northward shift and for range conservation, may be hampered by severe habitat fragmen-

tation in the more agricultural and urban landscapes of Europe. This negative interaction between landscape structure and climate warming has already been evidenced for habitat specialist butterfly species (Hill et al. 2001; Warren et al. 2001). It can be expected that the interaction between landscape structure and climate change is even more important for species groups that are less mobile than butterflies like, e.g. plant species. Here we studied the effects of landscape fragmentation on the colonization capacities of forest plant species. Most forest plant species are known to be specialist species adapted to the relatively stable environmental conditions of forests (Hermy et al. 1999). They usually have long generation times preventing rapid response to changing environmental conditions by in situ evolutionary adaptation (Bierzychudek 1982; Inghe & Tamm 1985). One of the ways to cope with the changing climate is to track the changing environment and to colonize new suitable forest fragments beyond their original northern range boundaries. In contrast to more mobile organisms like insects, range shifts for immobile species with long generations times have, as far as we know, not yet been observed and no time series are available. To cope with these methodological problems we Ó2002 Blackwell Science Ltd/CNRS

526 O. Honnay et al.

studied the migration potential of forest plant species in an indirect way and inferred a chronosequence from a toposequence. This means that because no time series was available, a snapshot of forests of different age was interpreted as a time series. We studied the colonization rates of a large number forest plant species by investigating the colonization success of the species from ancient source forest fragments into recently (c. 40 years) reforested target fragments. We distinguished between two degrees of fragmentation and considered (1) stepping stone dispersal in a highly fragmented forest landscape and (2) diffusionlike dispersal in a high connectivity forest landscape. METHODS

Stepping stone migration was studied in an agricultural landscape of 42 km2 where all 240 deciduous forest fragments (mean fragment area 1.1 ha, maximal fragment area 11.4 ha, total forested area 276 ha (6.6%), mean distance between target patch and source patch 210 m) were surveyed for the presence or absence of the 203 plant species which are confined to forest habitat in northern Belgium (for the species list and a more detailed description of the study area see Honnay et al. (1999) and Butaye et al. (2001), respectively). Forest fragments were defined as spatially segregated patches surrounded by non-forest land use. If there were differences in land use history between parcels within a patch, the patches were further subdivided in fragments originating before and after 1956. Most parcels originating before 1956 have a land use history of forest going at least back the oldest available topographical maps (c. 1775). Fragments were sampled twice: once in early spring and a second time during the end of the summer 1998. They were systematically surveyed by inspecting transects of 5 m wide through the whole of the patch and determining the presence of the 203 enlisted forest plant species. This resulted in a complete survey (1/0 data) of the complete perimeter of each forest patch. Eighty-four of the fragments (referred to as target fragments), were established after the year 1956, enabling us to determine the colonization success of the plant species from the older source fragments. For all plant species encountered in the target fragments we defined the colonization success as the number of occupied target fragments divided by the sum of the number of occupied target fragments and the number of suitable but not occupied target fragments. Habitat suitability of a target fragment was determined by creating a four-dimensional hypervolume for each source fragment based on the most extreme Ellenberg species indicator values for nitrogen, acidity, moisture and light for all plant species occurring in the forest (Ellenberg et al. 1992). A hypervolume is an abstract n-dimensional figure. A fragment is supposed to be Ó2002 Blackwell Science Ltd/CNRS

suitable for a plant species if the Ellenberg indicator values of that species fit in the hypervolume (Butaye et al. 2001). To gain insight into the relation between colonization success and distance between source fragment and target fragment we performed logistic regressions between the actual species presence/absence in a suitable target forest fragment and the average distance to the nearest five populations of that species in ancient source fragments. For statistical reasons the analysis was restricted to those species occurring in at least 5% of the target forest fragments. Diffusion-like migration was studied in a physically homogeneous forested landscape of 360 ha (total of 199 fragments, mean fragment area 0.6 ha, maximal fragment area 5.8 ha, total forested area 150 ha (42%)). The high number of fragments theoretically enables percolation of the species through the landscape. One-hundred and twentyfive of the forest fragments (referred to as target forest habitat) were established after 1956. The 199 fragments and the hedgerows in between were surveyed twice in 2000 for the occurrence of 12 forest plant species (Ranunculus ficaria, Geum urbanum, Festuca gigantea, Adoxa moschatellina, Paris quadrifolia, Lamium galeobdolon, Anemone nemorosa, Primula elatior, Arum maculatum, Deschampsia caespitosa, Listera ovata and Brachypodium sylvaticum). Plant species were selected to include both fast colonizers and slow colonizing species following Hermy et al. (1999) and Bossuyt et al. (1999). Hence there is no bias to be expected towards slow or fast colonizing species. In the forest parcels, the species were mapped by walking transects ± 10 m apart and recording every 10 m the presence/absence of the species. Outside the forests hedgerows were surveyed (all linear features composed of trees and/or shrubs) because these may be colonization sources. Again, the presence/absence of the species was recorded every 10 m. Afterwards the presence/ absence (p/a) of the species within the virtual grid cells was assigned to a forest parcel with a homogeneous land use history (i.e. the data were upscaled). The recording method may be slightly different between study areas; for both study areas we ultimately achieved perfectly comparable and accurate p/a data for each forest parcel. Colonization success of each species and the relation between colonization success and distance to the source populations was calculated as above. Soil analyses (soil texture, soil moisture and pH) indicated that all target forest habitats were suitable for the studied plant species, so we did not perform any corrections for differences in habitat quality. Because the age distribution of the forests in both of the study areas was comparable (most fragments were between 30 and 42 years old) we could compare the colonization success of forest plant species in the forest landscape with high connectivity vs. the more fragmented forest landscape. The colonization successes of the 12 species common in the

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two surveys were plotted against each other. Next, for the four forest plant species, common in the two surveys and exhibiting significant (Wald statistic, P < 0.05) logistic regression results (Geum urbanum, Primula elatior, Adoxa moschatellina and Arum maculatum) we plotted the estimated probability of occurrence of the species in a target habitat against the average distance to the nearest five source populations. We did this for the fragmented landscape and for the high connectivity landscape. Finally, in order to be able to predict the dispersal capacities of forest plant species, we associated the dispersal type of each species with its colonization success in the more fragmented landscape using a nonparametric ANOVA (Kruskal–Wallis test). Dispersal type was based on the presence of seed attributes like elaiosomes, awns and wings (cf. Hodgson et al. 1995). We distinguished between myrmecochore (dispersed by ants), barochore (no dispersal mechanism), endozoochore (dispersed by animals by ingestion), epizoochore (dispersed by animals through adhesion to, e.g. fur) or anemochore (dispersed by wind) dispersal.

areas’’) on the other side (Analysis of Covariance, F ¼ 7.32, P ¼ 0.013) (Fig. 2). The colonization success of the plant species in the high connectivity landscape was significantly higher than in the more fragmented landscape. The probability of occurrence of a species in the fragmented landscape at a certain distance from the source population was well below its probability of occurrence in the high connectivity landscape (Fig. 3). There was a marginally significant relation (Kruskal– Wallis test, Chi2 ¼ 9.16, d.f. ¼ 4, P ¼ 0.057) between 100% 80% 60% 40% 20% 0%

RESULTS

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Figure 1 Colonization success of forest plant species (1) in the

more fragmented landscape and (2) in the high connectivity landscape. Black, colonization success of less than 5%; shaded, exhibiting a negative occupancy vs. distance-to-source population relation; white, occurrence independent of distance from source populations.

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We found 112 forest plant species in the fragmented landscape. Thirty percent (34/112) of these species had a colonization success of zero, indicating that these species were not able to colonize suitable target forest fragments after 30–42 years. A cumulative total of 77% (86/112) of the plant species had a colonization success of less than 20%. For 20 of the species (18% of 112) we found a significant negative correlation between presence of the species and the distance measure (Wald statistic, P < 0.05). The remaining 17 species (15%) exhibited a distribution pattern that was independent of distance to possible source populations. These species show no dispersal limitation and they were able to colonize most suitable fragments throughout the study area (Fig. 1). In the high connectivity landscape two species (17%) had a colonization success of more than 81%. Eight species (66%) had a success rate lower than 20%. In a similar way as in the more fragmented landscape we performed a logistic regression for species that occurred in at least 5% of the target forest grid cells (50% or 6/12 species). Four species showed a significant relation with average distance to the five nearest source populations. Two species (17%) showed no significant relation and were able to migrate freely through the landscape and to colonize target patches independently of their isolation (Fig. 1). There was a significant difference in the slope between the regression line through the origin for the 12 species common in the two surveys on the one side and the y ¼ x regression line (the latter indicating ‘‘no difference in colonization success of the species between the two study

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Figure 2 Colonization success of 12 species common in the two

study areas. 1, Paris quadrifolia; 2, Brachypodium sylvaticum; 3, Deschampsia caespitosa; 4, Listera ovata; 5, Anemone nemorosa; 6, Lamium galeobdolon; 7, Festuca gigantea; 8, Primula elatior; 9, Arum maculatum; 10, Adoxa moschatellina; 11, Geum urbanum; 12, Ranunculus ficaria. Full line is the regression line through the origin (y ¼ 0.62x); dotted line is the y ¼ x line. Ó2002 Blackwell Science Ltd/CNRS


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Figure 3 Estimated probability of occurrence of a plant species in a target fragment as a function of the average distance to the five nearest

source populations of that species. Triangles, for the more fragmented landscape; filled circles, for the high connectivity landscape. a, Arum maculatum; b, Geum urbanum; c, Adoxa moschatellina; d, Primula elatior.

DISCUSSION

In the more fragmented forest landscape, colonization rates of 85% of the forest plant species to suitable but unoccupied fragments were not higher than a few metres/ year (Fig. 3). Thirty percent of the studied species were not even able to colonize any new suitable target habitat. In the high connectivity landscape the colonization rates were slightly higher but fell within the same order of magnitude. Because of the expected rapid climate change, needed migration rates to track the changing environment are expected to lie somewhere in the range between 3000 and 5000 m/y (Davis & Shaw 2001). It is known that the effective Holocene migration rates of forest plant species were higher than the ones expected from empirical plant colonization studies like this one (see Ó2002 Blackwell Science Ltd/CNRS

30 Colonization success ± 1 SE (%)

dispersal type and colonization success (Fig. 4). There is a significant trend for endozoochores and epizoochores to have a higher colonization success than the other dispersal types (Kruskal–Wallis pairwise comparisons, P < 0.05).

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Figure 4 Relation between species dispersal mode and coloniza-

tion success in the fragmented landscape. Ane, anemochore (dispersed by wind); bar, barochore (no dispersal adaptations); endo, endozoochore (dispersed by ingestion by birds and mammals); epi, epizoochore (dispersed by adherence to birds and mammals); myr, myrmecochore (dispersed by ants).

Forest fragmentation and plant migration 529

also Brunet & von Oheimb 1998; Bossuyt et al. 1999; Butaye et al. 2001). Since the last glacial most forest species were able to expand their range between 1000 and 2000 km (i.e. an average of hundreds of metres/year). This has been explained so far by the major influence of so-called chance events (i.e. a big importance is attributed to the tail of the seed number–distance distribution curve) (Cain et al. 1998; Higgins & Richardson 1999; Cain et al. 2000). Cain and colleagues could explain Holocene colonization rates of the North American forest herb Asarum canadense based on the tail of the seed dispersal curve of the species. The chance events associated with this tail may have been hurricanes or large migrating herbivores. The speed of climate warming combined with the degree of landscape fragmentation and the present absence of large migrating herbivores in the studied region will make the contribution of chance events in the dispersal and colonization process very minor. The fact that we found an association between colonization success of the studied species and a relatively rough measure for their dispersal ability is another indication for the absence of a major role for stochastic processes within the relatively short time span colonization happened. Our empirically determined colonization rates (metres/year) are two orders of magnitude lower than the required colonization rates, and one order of magnitude lower than the rates after the last glaciation. These rates are probably very realistic within the relatively short time period that climate warming is expected to occur. Even the immediate establishment of (continental) forest corridors or stepping stone forest fragments to enable northward migration (as suggested by e.g. Hunter et al. 1988) is hardly a solution, because of the low dispersal rates of the forest plant species. A lot of forest species have their northern range boundaries in Scandinavia. In some of these regions habitat fragmentation s.s. may be less of a problem. Habitat fragmentation from a forest plant species view is mainly a problem of forest management (nowadays dominated by spruce cultivation). On the other hand, seed dispersal by migrating herbivores is more likely in these regions. One could argue that the difficulties of the species in colonizing the target patches result from their early succession stage (and low habitat quality), and that the forest species which are present in the target patches were recruited from a persistent seed bank. The use of the Ellenberg Hypervolume, the fact that the forests are somewhere between 30 and 40 years old and that we indeed notice colonization (at least mostly in the less isolated patches), suggest that it is dispersal limitation and not recruitment limitation that hampers colonization of forest plant species. The view that (at least for forest plants) dispersal limitation is much more important than recruitment limitation is confirmed by more recent research (e.g. Brunet et al. 2000; Ehrlen & Eriksson 2000; Verheyen &

Hermy 2001). Regarding the presence of a seed bank, because the most important historical land use of target patches was grassland for more than 50 years, a remaining seed bank of forest species is nearly impossible, because these species are characterized by an extremely transient seed bank (Bossuyt & Hermy 2001; Bossuyt et al. 2002). The species that are the most sensitive to the consequences of global change in a fragmented landscape are those species which in an ecological restoration context are already known as so-called ancient forest plant species (Honnay et al. 2002). Ancient forest species are species which are confined to ancient forest fragments because they have serious difficulties in colonizing newly established forest patches, especially when these are spatially isolated. Hermy et al. (1999) give an overview of these species. They cover c. 30% of the forest flora sensu Ellenberg et al. 1992), a percentage which fits our 30% of species which are unable to leave their forest fragments. Given the actual landscape fragmentation and the short time period involved, we can conclude that colonization is very difficult for most forest plant species and that their survival will depend on their environmental tolerance. If their environmental tolerance turns out to be low they may be replaced by mobile generalist species invading from the south (see, e.g. McKinney & Lockwood 1999). We suggest much more research about environmental tolerance for the most immobile species (like the ancient forest species), based on, e.g. transplantation experiments. For many forest plant species transplantation or artificial seed transport to more northern forests may be the only solution to conserve their ranges, and possibly to save them from extinction. ACKNOWLEDGEMENTS

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Editor, N. G. Yoccoz Manuscript received 11 February 2001 First decision made 14 March 2002 Second decision made 5 April 2002 Manuscript accepted 16 April 2002