Forest fragmentation driven by demand for palm oil is having a catastrophic effect on multiple levels of biodiversity

Parallel declines in species and genetic diversity in tropical forest fragments

  1. Matthew J. Struebig1,2,*,
  2. Tigga Kingston3,
  3. Eric J. Petit4,
  4. Steven C. Le Comber1,
  5. Akbar Zubaid5,
  6. Adura Mohd-Adnan5,
  7. Stephen J. Rossiter1,*

Article first published online: 13 MAY 2011

DOI: 10.1111/j.1461-0248.2011.01623.x


Ecology Letters

Ecology Letters

Volume 14, Issue 6, pages 582–590, June 2011

Additional Information

How to Cite

Struebig, M. J., Kingston, T., Petit, E. J., Le Comber, S. C., Zubaid, A., Mohd-Adnan, A. and Rossiter, S. J. (2011), Parallel declines in species and genetic diversity in tropical forest fragments. Ecology Letters, 14: 582–590. doi: 10.1111/j.1461-0248.2011.01623.x

Author Information

  1. 1School of Biological and Chemical Sciences, Queen Mary University of London, London E1 4NS, UK
  2. 2Durrell Institute of Conservation and Ecology, School of Anthropology and Conservation, University of Kent, Canterbury CT2 7NR, UK
  3. 3Department of Biological Sciences, Texas Tech University, Lubbock, TX 79409-3131, USA
  4. 4University Rennes 1/CNRS, UMR 6553 ECOBIO, Station Biologique, F-35380 Paimpont, France
  5. 5Faculty of Science and Technology, Universiti Kebangsaan Malaysia, 43600 UKM Bangi, Selangor, Malaysia

*Correspondence: Matthew J. Struebig, Stephen J. Rossiter,

*Correspondence: E-mails:,;

Publication History

  1. Issue published online: 13 MAY 2011
  2. Article first published online: 13 MAY 2011
  3. Editor, Marcel Holyoak Manuscript received 9 February 2011 First decision made 8 March 2011 Manuscript accepted 28 March 2011


  1. Top of page
  2. Abstract
  3. Introduction
  4. Material and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  9. Supporting Information

Ecology Letters (2011) 14: 582–590


The potential for parallel impacts of habitat change on multiple biodiversity levels has important conservation implications. We report on the first empirical test of the ‘species–genetic diversity correlation’ across co-distributed taxa with contrasting ecological traits in the context of habitat fragmentation. In a rainforest landscape undergoing conversion to oil palm, we show that depauperate species richness in fragments is mirrored by concomitant declines in population genetic diversity in the taxon predicted to be most susceptible to fragmentation. This association, not seen in the other species, relates to fragment area rather than isolation. While highlighting the over-simplification of extrapolating across taxa, we show that fragmentation presents a double jeopardy for some species. For these, conserving genetic diversity at levels of pristine forest could require sites 15-fold larger than those needed to safeguard species numbers. Importantly, however, each fragment contributes to regional species richness, with larger ones tending to contain more species.


  1. Top of page
  2. Abstract
  3. Introduction
  4. Material and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  9. Supporting Information

Retaining habitat fragments as conservation set-asides within agricultural land is perceived as a valuable ‘wildlife-friendly’ management practice, and has received renewed interest following the recent accelerated conversion of rainforest to oil palm (Elaeis guineensis) plantations (Fitzherbert et al. 2008). Emerging eco-certification schemes require that environmentally responsible companies preserve areas of high conservation value within their concessions (Yaap et al. 2010). These areas, however, are typically small and disturbed, which calls into question their biodiversity value compared with large habitat areas outside the agricultural matrix (Edwards et al. 2010). If wildlife is to survive in tropical landscapes over the long-term, an understanding of how biodiversity is regulated by disturbance processes such as fragmentation is therefore required (Meijaard & Sheil 2007).

Processes that affect the diversity of species assemblages may also influence the genetic diversity, and hence long-term viability, of the populations within those assemblages. Co-variation in species and genetic similarity among sites can arise from random processes, which alter the composition of assemblages and genetic variants within them (Hubbell 2001). Additionally, classic island models of biogeography (MacArthur & Wilson 1967) and population genetics (Wright 1940) predict parallel declines in species and genetic diversity across islands and habitat fragments. Fragmentation results in increased assemblage and genetic drift via the loss of rare species and alleles respectively. Small fragments typically support fewer species and smaller populations than do larger fragments (Ewers & Didham 2006) and remaining species are susceptible to declines in genetic diversity and incur associated fitness costs (Vellend & Geber 2005; Keyghobadi 2007).

While the effects of drift can be offset by the immigration of new species, individuals and genes, this will depend on how well the surrounding habitat can facilitate or hinder dispersal (Ewers & Didham 2006; Keyghobadi 2007). Some have therefore argued for a link between the responses of species and genes to fragmentation (Vellend 2003) to the extent that one level of diversity might be used to predict the other (Cleary et al. 2006). If demonstrated, protecting habitats to maintain population genetic diversity as a co-benefit of conserving species diversity could further persuade land-managers to retain native habitat patches in modified landscapes for long-term biodiversity benefit.

Few studies have empirically tested for an association between species and genetic diversity in modified landscapes, but there is an emerging consensus that the responses of these two units to disturbance are linked by common processes. Cleary et al. (2006) demonstrated parallel declines in butterfly species and genetic diversity in response to forest fires. Similar associations have also been uncovered in temperate forest plants (Vellend 2004) and freshwater gastropods (Evanno et al. 2009), attributed to historical changes in land-use. To date, our main insights into species–genetic correlations in the context of fragmentation have come from post hoc comparisons of the results of community and genetic studies of islands or fragments, often undertaken at different times (for a review see Vellend 2003). The majority of island datasets reveal positive species–genetic diversity correlations (SGDCs) linked to area, a finding that has been confirmed in mainland habitat patches in recent years (Vellend 2004; He et al. 2008). However, results from other fragment datasets appear to be more equivocal (Vellend 2003). This weaker signal from fragments could reflect the shortcomings of experimental design, or alternatively, real differences in species’ responses to fragmentation that could arise with variation in both dispersal capabilities and the extent to which matrix habitats facilitate dispersal (Ewers & Didham 2006; Fischer & Lindenmayer 2007; Keyghobadi 2007).

Here, we report on the first empirical tests of an association between changes in species and genetic diversity across co-distributed taxa in the context of habitat fragmentation. We focus on insectivorous bats, which represent up to half of mammal species in Palaeotropical forests, and are known to experience major diversity declines in response to disturbance (e.g. Lane et al. 2006). Previous studies documented declines in bat diversity after forest conversion in our study region in line with trends reported for other vertebrates (Fitzherbert et al. 2008) and we have previously shown that insectivorous bat assemblages experience area-dependent losses in diversity and abundance in forest fragments (Struebig et al. 2008a).

Here, we compare trends in species and genetic diversity of insectivorous bats across a landscape undergoing major conversion to oil palm plantations in Southeast Asia. We examine associations within and between sites, and test the extent to which any such observed relationships are mediated by differences in habitat features such as fragment size and isolation. Additionally we consider the relative contribution of different fragments to landscape-wide diversity. Our results show that the magnitude of allelic loss varies among three species with different ecological traits, linked to their patterns of local dispersion and predicted capacity for movement. By comparing diversity (species and alleles) in fragments and undisturbed continuous habitat, we reveal the biodiversity savings associated with retaining fragments of various sizes. At the same time, we show that all fragments harbour unique elements of diversity and so contribute to overall regional levels.

Material and methods

  1. Top of page
  2. Abstract
  3. Introduction
  4. Material and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  9. Supporting Information

Study area

We undertook our research in central peninsular Malaysia (3°40′ N, 102°10′ E), a region historically covered with continuous dipterocarp rainforest, but now fragmented following industrial logging and plantation development over the last 40 years (Struebig et al. 2008a). The landscape matrix is dominated by plantations of oil palm and rubber (Hevea brasiliensis). Conversion of old rubber estates to oil palm is ongoing.

Assemblage and population genetic analyses were based on > 10 000 insectivorous bats captured between 2002 and 2007, with the vast majority of data obtained between 2005 and 2007. We sampled bats at 27 forest fragments (F01–F27) of different sizes and isolation histories, and five sites (S01–S05) in undisturbed continuous forest within the Krau Wildlife Reserve (Struebig et al. 2008a). National land-use maps and 2002 Landsat satellite images were used to identify sites representative of the landscape. Forest area was quantified using Arcview v.3.2 (ESRI, Redlands, CA, USA) and was log-transformed to approximate a normal distribution. For our indices of isolation, we generated two measures of geographic distance to the nearest forest fragment: first, we measured straight-line Euclidean distances, assuming a homogenous matrix with respect to animal movement. Second, we used least-cost modelling to calculate effective distances, to account for variation in landscape structure that could potentially impose differential resistance to movement. Each effective distance between sites represented the least-cost path that an animal would follow if matrix habitats impeded movement. To generate effective distances, the landscape was divided into a friction grid based on a land-use map in which each pixel represented c. 55 m. A least-cost algorithm was then used to determine the least-cost path between designated sites, describing the resistance to animal movement. The costs applied to the friction grid were high (50) for non-forest pixels, and low (1) for forest pixels to best represent a situation where an animal would favour forest areas and avoid non-forest (i.e. plantation habitats); narrow landscape elements (e.g. rivers, roads) would be crossed, but large non-forest areas avoided. As a result, effective isolation distances to the nearest forest patch were significantly longer than Euclidean (t = −4.127, P = 0.0006). Least-cost modelling was undertaken using the Pathmatrix extension to Arcview (Ray 2005).

Animal sampling

Bats were captured at each site using harp traps set across transects of comparable length (c. 1.0–1.5 km) that followed trails, streams or logging-skids. All trapping took place in dry seasons and was carefully standardized to avoid periods of heavy rain, which could influence capture success. Thus for assemblage-level analyses, sampling coverage and effort was standardized across sites in continuous and fragmented forest. The full sampling protocol is described in Kingston et al. (2006).

Tissue samples comprised two 3-mm diameter wing membrane biopsies (one from each wing) stored individually in 90–95% ethanol and were taken from adult bats using a biopsy punch (Stiefel Laboratories, Maidenhead, UK). We excluded recaptured individuals, which were recognized from biopsy scar tissue and, in some cases, by numbered forearm bands (Porzana, Icklesham, UK). All marked bats were released at the capture point following this procedure.

Study species for genetic analyses

Genetic analyses focused on three species that are relatively common in undisturbed forest: Blyth’s horseshoe bat (Rhinolophus lepidus), the trefoil horseshoe bat (Rhinolophus trifoliatus) and the papillose woolly bat (Kerivoula papillosa). These species exhibit contrasting ecological traits that are expected to result in differential responses to fragmentation. Rhinolophus lepidus is predicted to be the least dispersal-limited; it roosts in large colonies in caves and can forage far from the roost, dominating assemblages up to 11 km away (Struebig et al. 2009). As a likely result of this vagility, it exhibits no abundance response to fragmentation (Struebig et al. 2008a). In contrast, both other species roost in forest vegetation and show limited dispersal, characterized by small home ranges (typically < 100 ha around the roost) that do not typically extend beyond the forest edge (Kingston et al. 2006). However, whereas K. papillosa roosts in groups of 2–15 individuals in small tree cavities, R. trifoliatus is a solitary and more evenly distributed species, roosting alone in understory and midstory foliage (Kingston et al. 2006). Kerivoula papillosa also occurs at much lower densities than R. trifoliatus in undisturbed rainforest (Abdul-Aziz 2006). Such differences in local dispersion, population density and dispersal, all suggest that K. papillosa will be particularly vulnerable to stochastic loss and subsequent drift in fragments. Our previous data confirm that this species exhibits an area-dependent decline in abundance in forest fragments (Struebig et al. 2008a).

DNA extraction and amplification

We investigated genetic diversity using microsatellite markers, which are sensitive to fine-scale population variation and are one of the most powerful molecular tools used in conservation genetic research (Avise 2004). Genomic DNA was extracted from wing samples using Promega Wizard Purification Kits (Promega, Madison, WI, USA) and individual bats were genotyped at 8–15 unlinked microsatellite loci depending on the markers available for each species. For K. papillosa, we used 15 polymorphic loci (Kpa02, Kpa04, Kpa05, Kpa08, Kpa16, Kpa18, Kpa22, Kpa24, Kpa26, Kpa27, Kpa30, Kpa32, Kpa35, Kpa46 and Kpa47 – see Struebig et al. 2008b). For the two species of Rhinolophus, we used loci originally developed for other horseshoe bat species (Rossiter et al. 1999; Dawson et al. 2004; Puechmaille et al. 2005; E. Petit, unpublished work) that successfully cross-amplified in our study species (Table 1). All polymerase chain reactions were undertaken on a DNA Engine Tetrad Thermal Cycler (MJ Research, Waltham, MA, USA) using the procedure outlined in Struebig et al. (2008b). Alleles were then run on a 3700 sequencer (Applied Biosystems, Foster City, CA, USA), assigned sizes by Genescan (Applied Biosystems) and scored using Genotyper v3.6 (Applied Biosystems). In total, we genotyped 322 individuals of K. papillosa (99 from 11 fragments; 223 from continuous forest), 250 of R. trifoliatus (98 from 9 fragments; 152 from continuous forest) and 223 of R. lepidus (98 from 12 fragments; 125 from continuous forest) (Table 2).

Table 1.   Characterization of 14 microsatellite loci used for genetic analyses of Rhinolophus trifoliatus and Rhinolophus lepidus presented in Fig. 1d and e
Locus, accession number Trefoil horseshoe bat Blyth’s horseshoe bat
R. trifoliatus R. lepidus
41 individuals, site S05 35 individuals, site S02
TA (°C) MgCl2 (mm) Size range (bp) HO HE TA (°C) MgCl2 (mm) Size range (bp) HO HE
  1. Loci were originally isolated from other horseshoe bat species and tested in populations from large continuous habitat at optimum annealing temperatures (TA) and MgCl2 [for procedure see Struebig et al. (2008b)]. Heterozygosity (observed, HO; expected, HE) and deviations from Hardy-Weinberg equilibrium were assessed using a Markov-chain method implemented in Genepop v3 (Rousset 2008). None of the loci exhibited null alleles or consistent departure from Hardy-Weinberg expectations across the populations tested.

  2. *Primer sequences published in Rossiter et al. (1999).

  3. †Primer sequences published in Dawson et al. (2004).

  4. ‡Primer sequences published in Puechmaille et al. (2005).


Rferr08*, AF160207 56 3.0 140–186 0.750 0.911
Rferr11*, AF160210 54 3.0 172–196 0.765 0.862
Rferr14†, AJ560695 60 3.5 187–219 0.950 0.848 65 3.0 227–241 0.774 0.795
Rferr22†, AJ560704 53 3.0 168–200 0.857 0.880
Rferr27†, AJ560710 57 3.0 154–226 0.568 0.917 50 3.0 135–207 0.857 0.912
RHD107‡, DQ102694 51 3.0 214–242 0.853 0.793
RHA101§, JF750632 56 3.0 131–153 0.656 0.883
RHA104§, JF750633 50 3.0 264–296 0.829 0.791 55 2.0 281–316 0.806 0.910
RHA105§, JF750634 56 3.0 172–190 0.914 0.844
RHA107§, JF750635 51 3.0 127–159 0.900 0.878
RHA118§, JF750636 50 2.0 221–231 0.561 0.510 60 3.0 223–251 0.912 0.886
RHA7§, JF750630 50 3.0 222–258 0.825 0.862
RHA8§, JF750631 54 2.0 137–176 0.970 0.917
RHB112§, JF750637 55 2.0 187–225 0.974 0.918
Table 2.   The number of individuals genotyped (N), abundance (at standardized sample effort –n) mean alleles per locus (A) and rarefied allelic richness (RS) for populations of three bat species sampled in forest fragments (F01–F24) and continuous forest that are presented in Fig. 1c–e
Site Area (ha) Euclidean distance (km) Effective distance (km) Papillose woolly bat Trefoil horseshoe bat Blyth’s horseshoe bat
Kerivoula papillosa Rhinolophus trifoliatus Rhinolophus lepidus
(15 loci) (8 loci) (10 loci)
N n A RS N n A RS N n A RS
  1. Area and isolation distances of forest sites are also included. Allelic richness data from the continuous forest are presented as mean values across five replicate sites (S01–S05) and were used to produce the reference lines in Fig. 1c–e.

F01 2883 2.29 3.32 14 2 6.600 6.761 14 12 8.750 8.983 2
F02 11 339 4.32 5.32 15 8 6.467 6.467 14 8 8.750 8.956 10 11 9.000 9.503
F03 100 3.53 3.60 0 0 25 16 12.400 9.697
F04 551 2.10 4.85 3 8 5 7.675 10.914 0
F06 443 2.24 2.97 11 8 5.600 6.097 4 12 10 9.200 9.200
F07 1356 2.45 2.52 11 2 5.667 6.276 0 18 210 10.300 9.643
F08 353 1.80 3.51 4 4 3.067 3.269 6 6 5.500 7.305 0
F09 31 0.9 1.02 1 4 6 7 6.900 9.041
F12 319 2.44 2.86 1 6 2 5.625 7.711 21 13 11.300 9.793
F14 400 3.55 4.46 4 3 3.533 4.995 9 8 7.125 8.000 2
F15 251 2.29 2.73 1 2 13 16 10.100 10.100
F16 93 1.03 1.19 0 0 4 4 5.900 9.596
F17 32 1.09 1.14 0 10 6 8.875 10.918 9 7 7.700 9.546
F18 115 0.73 1.61 8 9 4.867 5.675 0 6 7 6.800 9.830
F19 100 1.44 1.67 0 0 8 8 8.600 9.723
F20 300 0.75 0.89 0 0 18 43 10.200 9.662
F21 5225 4.67 4.86 10 5 5.267 6.177 21 9 10.375 9.426 0
F22 102 3.40 3.71 5 5 3.600 5.204 10 5 7.000 8.070 0
F23 5581 2.67 4.58 11 11 5.467 6.119 3 2
F24 2025 2.64 2.94 6 6 4.467 5.065 3 40
Cont. 137 000 1.1 1.1 223 45 (7–11) 7.813 6.332 152 21 (2–8) 11.511 9.474 125 27 (0–12) 10.300 9.531

Analysis of alpha diversity within sites

To quantify site-level diversity, we chose richness-based measures, which should be more sensitive to the elimination of rare species and alleles caused by drift (Allendorf 1986), as well as being intuitively the easiest of measures to interpret. As diversity measurement is highly dependent on sample size, we estimated species richness of all insectivorous bats at each site at a standard trapping effort (15 harp traps) using sample-based rarefaction (1000 randomizations) in EstimateS v7 ( We also determined separate richness values for forest interior specialists that are characterized by roosting preferences for tree cavities and foliage (Struebig et al. 2008a). Rarefaction was also used to compare levels of allelic richness, a genetic analogue of species richness that serves as a direct observation of the number of alleles within a population. We used the ARES package (van Loon et al. 2007) of R v2.5.1 (, to rarefy allelic richness to a common sample size of 15 individuals. For sites where fewer than 15 individuals were captured, the ARES algorithm was used to extrapolate mean allelic richness values up to this sample size, using 200 bootstrap re-samples to generate confidence intervals. This method, akin to that used for estimating richness from species accumulation curves, performs well when extrapolating to estimates before the curve asymptotes (Colwell et al. 2004), which was the case for all populations in this study. To compare fragment site-wise estimates of species and allelic richness to levels seen in intact habitat, we also derived mean values, and associated confidence limits, from five replicate sites within continuous rainforest.

To determine whether variance in species and genetic diversity among fragments was driven by fragment area or isolation, we fitted separate general linear models (GLMs) for each species, in R. Data for genetic diversity models are reported here for the first time. However, standardized species data from 15 fragments were reported in Struebig et al. (2008a). In addition to our GLMs, we also used partial (Pearson) correlations to test the relative importance of area and isolation in determining diversity indices. Using this approach, we were also able to test whether any observed correlation between species and genetic diversity (i.e. the SGDC) remained after correcting for the effect of these landscape metrics and population size (using the standardized abundance of a species in 15 harp traps as a proxy).

Analysis of beta diversity between sites

To test for an association between species diversity and genetic diversity among sites, we compared matrices of pairwise assemblage dissimilarity and genetic differentiation distances. We used the Morisita-Horn index, calculated in SPADE (, to measure assemblage dissimilarity. This index is robust and most sensitive to differences in dominant species, and hence suited to situations in which rare species may be missing in small inventories (Chao et al. 2008). Genetic differentiation was quantified using the Jost D estimator, a genetic analogue of the Morisita-Horn index based on allelic richness, which measures allele fidelity to demes (Chao et al. 2008; Jost 2008). We also quantified differentiation using the more traditional Weir-Cockerham estimate of Fst, calculated in Genepop v3 (Rousset 2008). Mantel tests were used to determine if correlations between these diversity indices were statistically significant. As assemblages and populations are also expected to be more differentiated with increasing geographic distance (Hubbell 2001), we also used this method to test for an association between each of the diversity indices and both Euclidean and effective geographical distance, predicting a stronger relationship with the latter (Broquet et al. 2006). All Mantel tests were undertaken in Genalex v6.3 (Peakall & Smouse 2006) with P values generated using 999 random permutations.

Cumulative regional (gamma) diversity

As similar species richness values across fragments can be based on either the same or different subsets of taxa present, we also examined cumulative diversity for combinations of fragments. This approach allowed us to compare the contribution, and thus importance, of multiple fragments for safeguarding regional diversity. We developed a script in Mathematica 5.0 (Wolfram Research, Champaign, IL, USA) to calculate species accumulation estimates by combining site-wise data from all 15 fragments in a random order. This was repeated 10 000 times, at each step retaining information on cumulative area and the number of fragments. To determine the extent to which regional-level species diversity is represented within sets of small or large fragments, we also plotted cumulative diversity by adding sites together in order of ascending and descending area respectively. To compare overall regional species diversity between fragments vs. continuous sites, we pooled all capture data within each dataset and rarefied to a common sample size. This analysis was also repeated for allele data. Finally, all randomization and rarefaction tests of species diversity were repeated for forest specialists only.


  1. Top of page
  2. Abstract
  3. Introduction
  4. Material and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  9. Supporting Information

We captured an average of 13.6 (CI 95%: 11.6–15.6) species in the five continuous forest sites based on a standardized sampling effort of 15 harp traps. For forest specialists 7.6 (CI: 6.6–8.6) species were captured. General linear models revealed that patch area, as opposed to either measure of isolation, was the best predictor of bat species richness declines in fragments (all species: adjusted r2 = 0.31, P = 0.018; forest specialists: r2 = 0.34, P = 0.010). Considering all species, richness in larger fragments (> 150 ha) typically fell within the confidence bounds of that of undisturbed continuous rainforest, with diversity levels in fragments of > 650 ha showing equivalent levels to the latter (Fig. 1a). Fewer species occurred in smaller fragments, while the largest of fragments exhibited elevated species richness due to higher numbers of generalist, edge-tolerant species. The resulting species–area relationship was not an artefact of this elevated richness in the largest fragments: when the five large fragments were excluded, correlation remained strong (adjusted r2 = 0.41, P = 0.028), whereas when the small fragments with depressed richness were removed, no correlation was evident (r2 = −0.11, P = 0.759). Furthermore, when considering only the subset of forest specialist species in the analyses, much larger fragments (> 2500 ha) were needed to reach a diversity level equivalent to that of continuous forest (Fig. 1b).

Figure 1.   The impact of rainforest fragmentation on species diversity (a,b) and the genetic diversity of populations of three tropical bat species (c–e). Species are presented in decreasing order of predicted sensitivity to fragmentation based on ecological traits. Estimates of the number of species and alleles in fragments of various sizes account for variation in sampling effort via rarefaction and are presented as a proportion of levels in continuous undisturbed rainforest. Dashed horizontal lines indicate the 95% confidence limits (CLs) of richness levels in this control habitat. Thus, values above the upper 95% CL represent significantly higher richness than in continuous forest, while values below the lower 95% CL represent significantly lower richness. Dotted lines denote the 95% confidence envelope of the line of best fit, which are presented where general linear models (GLMs) with area as a predictor are significant. See also Tables 1 and 2.


The impact of fragment size on the magnitude of loss of genetic diversity differed markedly across the three species studied. The species considered to be most sensitive to fragmentation, K. papillosa, exhibited a decline in allelic richness in fragments that correlated with fragment area (r2 = 0.39, P = 0.024, GLM, see Fig. 1c). By comparison, univariate GLMs showed that neither measures of isolation, nor abundance, had significant effects on diversity and these potential explanatory variables were absent from our best fitting models based on Akaike Information Criteria (AIC) (data not shown). Typically, large fragments of > 5000 ha showed genetic diversity levels that fell within the lower confidence interval of continuous forest.

Reduced allelic richness in fragments relative to continuous forest was also observed in most populations of R. trifoliatus (see Fig. 1d). Here, variance in richness was not explained by isolation; AIC could not discriminate area from abundance as the best predictor, and no overall area-dependent decline was evident (r2 = 0.34, P = 0.06, GLM). In the species predicted to be least dispersal-limited, R. lepidus, we found that allelic richness was similar across the vast majority of fragments and continuous forest (Fig. 1e), indicating no impact of fragmentation on genetic diversity. For this species, although AIC ranked effective isolation distance as the best predictor of variation in allelic richness in fragment populations, the overall model was not significant (r2 = 0.19, P = 0.09). For all three species, our GLMs were supported by the results of partial correlation tests (Table 3). In summary, we found that fragment area was the only determinant of allelic richness for K. papillosa, and we found no significant effect of either isolation or abundance in any of the species studied. From these results, we can also conclude that much more forest would be needed to maintain genetic diversity in K. papillosa than overall levels of species richness. Fragments would typically need to be > 10 000 ha in size to support populations with equivalent (i.e. mean) genetic diversity than in continuous rainforest (Fig. 1c).

Table 3.   Pearson correlation and partial correlation coefficients between allelic richness and area, and species richness and allelic richness, while controlling for the effect of other variables (shown in parentheses)
Species Kerivoula papillosa Rhinolophus trifoliatus Rhinolophus lepidus
  1. ‘Isolation’ refers to the effective geographic distance to the nearest forest patch. The same conclusions arise when using Euclidean distances.

  2. *Significant: P < 0.05.

Allelic richness × area 0.671 (0.024)* 0.455 (0.219) 0.347 (0.269)
Allelic richness × area (isolation) 0.777 (0.001)* 0.482 (0.178) −0.083 (0.803)
Allelic richness × area (abundance) 0.665 (0.012)* 0.122 (0.763) 0.268 (0.403)
Species richness (all bats) × allelic richness 0.578 (0.081) 0.398 (0.289) −0.023 (0.958)
Species richness (all bats) × allelic richness (area) 0.384 (0.271) 0.292 (0.455) −0.193 (0.661)
Species richness (forest bats) × allelic richness 0.489 (0.158) −0.198 (0.611) 0.021 (0.961)
Species richness (forest bats) × allelic richness (area) 0.505 (0.121) −0.199 (0.618) −0.037 (0.934)

Interestingly, despite the clear parallel species–genetic impacts observed in K. papillosa and a substantial correlation coefficient between these responses, association between the two levels of diversity was not significant (Pearson r = 0.578, P = 0.080; Table 3). Moreover, when controlling for the known influence of fragment area on diversity using partial correlations, the correlation dropped in magnitude and remained non-significant (r = 0.384, P = 0.271), suggesting that any SGDC was mediated by differences in fragment size.

Our analyses of beta diversity between fragments also confirmed that geographical isolation had no detectable impact on either species or genetic diversity. Assemblage dissimilarity varied substantially between pairs of sites in continuous forest (range of Morisita-Horn coefficients = 0.161–0.789) and also between fragments (coefficients for pairwise distances equivalent to those in continuous forest = 0.022–0.900), but Mantel tests revealed that assemblage dissimilarity between pairs of fragments was not correlated with Euclidean (rM = 0.011, P = 0.157) or effective isolation distance (rM = 0.004, P = 0.233). Values of D and Fst for continuous forest populations were generally low, indicating minimal genetic differentiation for any of the three study species (K. papillosa, D = 0.003–0.052, Fst = 0.006–0.023; R. trifoliatus, D = 0.000–0.050, Fst = 0.000–0.012; R. lepidus, D = 0.000–0.027, Fst = 0.006–0.023). No significant genetic isolation by distance was detected for any of the study species among fragments and there was poor concordance between assemblage dissimilarity and genetic differentiation using either Jost’s D (Table 4) or Fst (see Table S1).

Table 4.   Mantel test correlation coefficients between genetic differentiation measured by D (Jost 2008) and the following pairwise ecological or geographic distances for the three bat species: (i) bat assemblage dissimilarity (Morita-Horn index), (ii) Euclidean geographic distance and (iii) effective (i.e. least-cost) distance
Species Kerivoula papillosa Rhinolophus trifoliatus Rhinolophus lepidus
  1. Mantel tests were calculated for these pairwise distances in Genalex v6.3 (Peakall & Smouse 2006). Probabilities of the correlation coefficients, calculated using 999 permutations, are shown in parentheses.

(i) Assemblage dissimilarity (all bats) 0.0013 (0.485) 0.0257 (0.085) 0.1149 (0.085)
Assemblage dissimilarity (forest bats) 0.0015 (0.327) 0.1911 (0.058) 0.1772 (0.095)
(ii) Euclidean isolation distance 0.0085 (0.294) 0.0039 (0.287) 0.0299 (0.287)
(iii) Effective isolation distance 0.0222 (0.195) 0.0038 (0.444) 0.0046 (0.444)

Randomization analyses undertaken to assess the consequences of combining multiple fragments for species richness showed that diversity can vary widely for either a given cumulative area (Fig. 2a) or number of fragments (Fig. 2b) and that similar habitat areas can be represented by a range of fragment numbers (Fig. 2c). Inspection of species accumulation curves based on fragments of ascending and descending size revealed that, for a given cumulative area, the total number of species contained within several small fragments nearly always exceeded those of fewer larger fragments (Fig. 2a). However, for a particular number of fragments, larger sites consistently contained more species than did smaller ones (Fig. 2b). These trends also held for just forest specialists (see Figure S1).

Figure 2.   The contribution of multiple fragment combinations to regional species richness. Species accumulation was calculated by combining site-wise data from 15 fragments in a random order, while retaining information on the number of fragments, species and cumulative area. Species richness estimates are shown for 10 000 random fragment combinations (small grey points) and for fragments added together in order of ascending (i.e. smallest first, large red/black outline) or descending (i.e. largest first, large blue points) area. Data are presented as plots of cumulative forest area (a), cumulative number of fragments (b) and both (c). Data for forest specialists are presented separately in Figure S1.


Rarefied estimates of total regional diversity from pooled data suggested more species in fragments than continuous forest (common n = 404 bats), possibly due to edge-tolerant generalists (see Figure S2a). Indeed, no difference was evident based on forest interior specialists only (common n = 171; Figure S2b). For our genetic data, computational difficulties of retaining allele identity when combining individual-wise diversity values for unequal sample sizes precluded the use of our randomization procedure. However, rarefied estimates of regional allelic richness from pooled data suggested that, when compared with continuous data, fragments contained more alleles in K. papillosa (common n = 99 bats; Figure S2c) and R. lepidus (common n = 98; Figure S2e), but that levels were similar in R. trifoliatus (common n = 98; Figure S2d). Elevated numbers of alleles might reflect genetic drift in the least mobile species due to population subdivision, or could stem from the neutral variation across the region given that the maximum distance between continuous forest sites was less than that between fragments.


  1. Top of page
  2. Abstract
  3. Introduction
  4. Material and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  9. Supporting Information

Few empirical studies have set out to compare community and genetic responses in the context of habitat fragmentation, and to our knowledge, this is the first attempt to apply this approach to co-distributed species. Examining three co-distributed bat species, we find evidence that area-dependent declines in species richness across a fragmented landscape are mirrored by parallel declines in genetic diversity in some, but not all, species. Greatest allelic losses within fragments were observed in K. papillosa leading to a genetic-area relationship. This is a colonial tree-roosting species that occurs at low population densities and exhibits both a clumped local dispersion and limited capacity for dispersal. In contrast, in R. lepidus– a species capable of commuting over large distances between fragments – we observed no site-wise loss of allelic richness in fragments. In R. trifoliatus, a solitary and more evenly dispersed species, we detected allelic losses in most fragments, but no genetic-area relationship. Consequently, while our results from one species add to previous reports of the so-called SGDC, we suggest that such trends are likely to heavily depend on the ecological traits of the focal taxa.

Despite increased ecological genetic research in recent years, there is little consensus on the species and genetic consequences of fragmentation. In a review of 45 comparisons of fragmented vs. continuous habitats, only 58% reported reduced genetic diversity in fragments (Keyghobadi 2007). Such discrepancies may partially arise from the choice of molecular markers. For example, mitochondrial DNA remains popular in conservation genetics despite its relatively low resolution – a factor that might have reduced the power to detect genetic erosion in fragment populations of tropical bats (Meyer et al. 2008) and butterflies (Benedick et al. 2007). Similarly, published studies have used a range of diversity indices, which vary in their ability to detect population genetic responses to habitat change (e.g. Stow & Briscoe 2005). Our study benefits from both the use of microsatellite loci, which are able to detect fine-scale genetic variation, and richness-based measures that are analogous for the genetic and assemblage components of alpha and beta diversity under study.

We believe that our inability to detect an effect of fragment area on the genetic diversity of all species under study likely reflects ecological differences between these taxa, which will in turn influence the time lag in the genetic response to changing landscape structure (Anderson et al. 2010). Delays in the manifestation of fragmentation effects (i.e. extinction debts) are attributed to population crowding in fragments and uneven fragmentation histories (Ewers & Didham 2006). From a population genetic perspective, longer lags are expected in taxa with large effective population sizes (Excoffier 2004), long generation times (Kramer et al. 2008) and higher dispersal rates among patches. While unlikely to vary substantially in their generation times, our three study species vary in their capacities for movement. Although we found no impact of fragment isolation on diversity, fragmentation is a relatively young process in our study area and old plantations can be structurally similar to forest (at least relative to non-tree plantations). Substantial delays might therefore be expected before the effects of isolation and drift are detected on fragment populations and assemblages. Overall, our findings suggest that the observed negative impacts of fragmentation on bat species and genetic diversity in our study landscape have arisen via the passive sampling of individuals and/or species, and subsequent drift effects, both of which stem from reducing habitat area. However, the long-term role of isolation in shaping species and genetic diversity remains unknown.

Variation in species’ responses to habitat fragmentation is a theme common to both ecological genetics and community ecology. Previous studies suggest that species sensitivity to fragmentation is determined by traits linked to mobility, niche breadth and exploitation of matrix habitats (Ewers & Didham 2006; Fischer & Lindenmayer 2007; Keyghobadi 2007). Comparative approaches such as ours could help to identify vulnerable groups in diverse assemblages in intact habitats before they are exposed to land-use change (e.g. Kraaijeveld-Smit et al. 2007). Although consideration of additional species as well as information on other genetic markers and/or quantitative traits are needed to verify our findings, we stress that observed genetic losses associated with fragmentation in our study are unlikely restricted to a single species. K. papillosa shares a common ecological syndrome with > 50 phylogenetically related bat species (including 20 congeners) in tropical rainforest, all of which roost in trees, and possess wing morphologies and echolocation signals that enable efficient foraging in dense forest, but which constrain their use of more open habitats (Kingston et al. 2006). More generally, such sensitivity to disturbance events can be seen in subsets of most taxonomic groups (Fischer & Lindenmayer 2007).

An association between patch area and both allelic diversity in K. papillosa and overall species richness, supports previous suggestions that SGDCs (Vellend 2004; He et al. 2008; Odat et al. 2010) are mediated by habitat features. A substantial but non-significant SGDC in our study points to complex processes shaping different units of diversity. Increased sampling might add more statistical power; however, several smaller fragments have since been cleared for plantations. Correlations can also be weak or non-significant where one level of diversity is more sensitive to change than the other (Vellend 2005). Thus, while proposals to predict the environmental response of diversity components from each other is desirable for conservation (Cleary et al. 2006) caution must be exercised to avoid oversimplification, which could risk overlooking more vulnerable species.

Our findings have several important conservation implications. Comparisons of diversity levels in single fragments relative to those of intact habitat suggest that larger patch sizes are needed for maintaining similar levels of genetic diversity than for maintaining equivalent species diversity. Species diversity levels typical of intact forest were found in fragments of 650 ha when all species were considered and 2500 ha when focussing on forest specialists. Equivalent levels of genetic diversity in K. papillosa needed at least 10 000 ha, far exceeding previous estimates of 4000 ha for insects on Borneo (Bickel et al. 2006), and greater than most forest blocks remaining in production areas in Southeast Asia. However, considering fragments as independent units not only oversimplifies the bigger picture but also represents a situation unfamiliar to landscape managers who are typically tasked with choosing multiple patches to make up a conservation set-aside of predetermined area. We found that by combining data from multiple fragments, cumulative species diversity dramatically increased, even where the total area was relatively small. Thus, most or all fragments contributed to regional diversity levels. At the same time, for a given number of fragments, maximizing cumulative area always supported more taxa.

Despite higher overall species diversity in groups of small fragments, it is these sites that are more susceptible to delayed fragmentation effects and the impacts of stochastic events (Ewers & Didham, 2006). We therefore advocate calls to direct conservation investments in production landscapes towards well-managed large patches (Edwards et al. 2010). However, our analyses confirm the value of existing small habitat fragments, which are currently favoured by certification policy in agricultural estates (Yaap et al. 2010). Therefore, heeding lessons from the Single Large or Several Small (SLOSS; see Ewers & Didham, 2006) debate, we stress that these two strategies should not be viewed as mutually exclusive. Specifically, we note that in many regions worthy of conservation efforts, there are few, if any, large areas of intact habitat within the vicinity of agriculture. Here, conserving a network of patches can capture most of the regional diversity, however, landscape management that favours connectivity could help to maintain long-term diversity by encouraging dispersal and maximizing habitat area (Gillies & St. Clair 2008; Gilbert-Norton et al. 2010). This approach could also increase genetic diversity levels in isolated populations (Hale et al. 2001). Mitigation investments for commodity products such as palm oil should therefore be part of a landscape-wide conservation strategy in which small habitat fragments are conserved and linked as corridors to maximize habitat area and promote connectivity among large core areas of forest.


  1. Top of page
  2. Abstract
  3. Introduction
  4. Material and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  9. Supporting Information

This work was supported by grants awarded to MJS and SJR from the UK Natural Environment Research Council (NERC), Bat Conservation International/US Forest Service and the University of London Central Research Fund; to TK from the US National Science Foundation (NSF # 0108384, DEB & East Asia and Pacific Program), Earthwatch Institute, National Geographic (Committee for Research & Exploration; Conservation Trust) and Lubee Bat Conservancy; and from the Brittany Region to EJP. MJS was also funded by a Leverhulme Trust Early Career Fellowship, and SJR by a Royal Society Research Fellowship. We are grateful to Sheema Abdul-Aziz for sharing her population estimates of our study species and to three anonymous referees whose suggestions greatly improved our manuscript. We thank the Economic Planning Unit, Department of Wildlife and National Parks, Pahang State Forestry Department and the Federal Land Development Authority for permission to undertake our research in Malaysia.


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  2. Abstract
  3. Introduction
  4. Material and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  9. Supporting Information
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Supporting Information

  1. Top of page
  2. Abstract
  3. Introduction
  4. Material and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  9. Supporting Information

Figure S1 Contribution of multiple fragment combinations to regional species richness of forest specialist bats.

Figure S2 Regional rarefaction curves for species and alleles of the three study species in continuous forest and fragments.

Table S1 Mantel test correlation coefficients between genetic differentiation measured by Fst/1−Fst and ecological or geographic distances for the three study species.

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