Long-Term Dynamics of a Fragmented Rainforest Mammal Assemblage

Contributed Paper Long-Term Dynamics of a Fragmented Rainforest Mammal Assemblage WILLIAM F. LAURANCE,∗ ‡ SUSAN G. LAURANCE,∗ AND DAVID W. HILBERT† ∗ Smithsonian Tropical Research Institute, Apartado 0843-03092, Balboa, Ancón, Panama †CSIRO Sustainable Ecosystems, Tropical Forest Research Centre, P.O. Box 780, Atherton, Queensland 4883, Australia Abstract: Habitat fragmentation is a severe threat to tropical biotas, but its long-term effects are poorly understood. We evaluated longer-term changes in the abundance of larger (>1 kg) mammals in fragmented and intact rainforest and in riparian “corridors” in tropical Queensland, with data from 190 spotlighting surveys conducted in 1986–1987 and 2006–2007. In 1986–1987 when most fragments were already 20–50 years old, mammal assemblages differed markedly between fragmented and intact forest. Most vulnerable were lemuroid ringtail possums (Hemibelideus lemuroides), followed by Lumholtz’s tree-kangaroos (Dendrolagus lumholtzi) and Herbert River ringtail possums (Pseudocheirus herbertensis). Further changes were evident 20 years later. Mammal species richness fell significantly in fragments, and the abundances of 4 species, coppery brushtail possums (Trichosurus vulpecula johnstoni), green ringtail possums (Pseudochirops archeri), red-legged pademelons (Thylogale stigmatica), and tree-kangaroos, declined significantly. The most surprising finding was that the lemuroid ringtail, a strict rainforest specialist, apparently recolonized one fragment, despite a 99.98% decrease in abundance in fragments and corridors. A combination of factors, including longterm fragmentation effects, shifts in the surrounding matrix vegetation, and recurring cyclone disturbances, appear to underlie these dynamic changes in mammal assemblages. Keywords: Australia, cyclones, forest fragmentation, long-term research, mammal assemblages, marsupials, matrix vegetation, Queensland, tropical rainforests Dinámica a Largo Plazo de un Ensamble de Mamı́feros de un Bosque Lluvioso Fragmentado Resumen: La fragmentación del hábitat es una severa amenaza para las biotas tropicales, pero se conoce poco sobre sus efectos a largo plazo. Evaluamos cambios de largo plazo en la abundancia de mamı́feros mayores (>1 kg) en bosque lluvioso fragmentado e intacto y en “corredores” ribereños en Queensland, con datos de 190 muestreos con reflector llevados a cabo en 1986–1987 y 2006–2007. En 1986–1987, cuando la mayorı́a de los fragmentos ya tenı́an entre 20 y 50 años, los ensambles de mamı́feros difirieron significativamente entre bosque intacto y fragmentado. Los más vulnerables fueron zarigüeyas (Hemibelideus lemuroides), seguidas por canguros arborı́colas (Dendrolagus lumholtzi) y zarigüeyas (Pseudocheirus herbertensis). Mayores cambios fueron evidentes 20 años después. La riqueza de mamı́feros decayó significativamente en los fragmentos, las abundancias de 4 especies, Trichosurus vulpecula johnstoni, Pseudochirops archeri, Thylogale stigmatica y D. lumholtzi, disminuyó significativamente. El hallazgo más sorprendente fue que H. lemuroids, un especialista de bosque estricto, aparentemente recolonizó un fragmento, no obstante un 99.98% de disminución en abundancia en los fragmentos y corredores. Una combinación de factores, incluyendo los efectos de la fragmentación a largo plazo, cambios en la matriz de vegetación circundante y las perturbaciones ciclónicas recurrentes, parece subyacer en estos cambios dinámicos en los ensambles de mamı́feros. Palabras Clave: Australia, bosques lluviosos tropicales, ciclones, ensambles de mamı́feros, fragmentación de bosque, investigación a largo plazo, marsupiales, matriz de vegetación, Queensland ‡email [email protected] Paper submitted November 30, 2007; revised manuscript accepted January 30, 2008. 1154 Conservation Biology, Volume 22, No. 5, 1154–1164 C 2008 Society for Conservation Biology. No claim to original US government works. Journal compilation  DOI: 10.1111/j.1523-1739.2008.00981.x 1155 Laurance et al. Introduction Forest loss and fragmentation are among the most serious of all perils to tropical biodiversity (Laurance & Bierregaard 1997; Lindenmayer & Fischer 2006). Not only are tropical landscapes being rapidly fragmented, but many protected areas are becoming isolated as their surrounding habitats are degraded or destroyed (Curran et al. 2004; DeFries et al. 2005; Mayaux et al. 2005). Tropical forest disruption is expected to continue apace over the next century, driven by continued population growth, rapid industrialization, and agricultural expansion for crops and biofuels (Tilman et al. 2001; Rudel 2005; Laurance 2007). Clearly, the fate of much of tropical biodiversity will depend on our capacity to understand and limit the long-term impacts of forest fragmentation. Much is unknown about the dynamics of fragmented tropical ecosystems, especially over longer timescales. Some of the clearest insights have come from an experimentally fragmented landscape in Amazonia (Lovejoy et al. 1986; Laurance et al. 2002; Ferraz et al. 2003; Stouffer et al. 2006) and from small islands in a Venezuelan reservoir (Terborgh et al. 2001; Feeley & Terborgh 2006), but fragments in these investigations were just 1–2 decades old. Only in a handful of studies have researchers examined older fragmented systems, such as centuries-old forest remnants in Singapore and Hong Kong (Corlett & Turner 1997; Brook et al. 2003) and Pleistocene land-bridge islands (Terborgh 1975; Wilcox 1978). Here we provide a 20-year comparison of the abundances and species richness of tree-kangaroos, wallabies, and larger (>1 kg) possums in fragmented and intact rainforests in tropical Queensland, Australia. Our study area, the southern Atherton Tableland, is a key center of endemism that supports numerous mammal species with highly restricted geographical and elevational ranges (Winter 1988; Williams & Pearson 1997; Kanowski et al. 2001). We initially assessed the impacts of forest fragmentation on this fauna in 1986–1987 (Laurance 1990, 1991a, 1994, 1997), when most of the fragments were 2–5 decades old. We returned exactly 20 years later, to evaluate subsequent changes in mammal populations. This comparison provided rare insights into the longer-term dynamics of a fragmented species assemblage. Methods Study Area The Atherton Tableland (600–900 m elevation) in tropical Queensland was formerly dominated by rainfor- est. This cloudy, wet area (mean rainfall approximately 2800 mm/year on the southern Tableland) is an apparent Pleistocene refugium and is considered the most important center of species endemism in tropical Australia (Winter 1988; Williams & Pearson 1997). Much of the Tableland and has been deforested, mostly for dairy farming, cattle ranching, and residential development. Clearing began about 1909 and proceeded rapidly for the next 3 decades. By 1983 more than 76,000 ha of forest had been removed (Winter et al. 1987), leaving more than 100 rainforest fragments ranging from 1 to 600 ha in area, scattered over an area of about 900 km2 . Large (>3600 ha) tracts of intact but selectively logged rainforest, protected since 1988 as a World Heritage area, persist on steeper hillsides that enclose the margins of the Tableland. Fragments are surrounded by mosaics of cattle pastures and narrow (10–50 m wide) strips of forest regrowth along streams (hereafter termed corridors). The southern half of the Atherton Tableland (Fig. 1), where this study was conducted, has experienced only limited change in forest cover over the past 3 decades (determined on the basis of aerial imagery taken in 1986 and 2006). Recent forest clearing has been minimal, and the forest has partially regenerated in certain areas, most notably around a small (12.8-ha) fragment (fragment 5, Fig. 1) 220 m from intact forest. The largest disturbances have been caused by cyclones. The study area was damaged by major cyclones in 1986 and 2006 (Laurance 1991b; Laurance & Curran 2008), in each case just 5– 7 months before our spotlighting surveys commenced. The region also suffered strong El Niño related droughts in 1982 and 2002 that caused substantial animal and plant mortality (N. I. J. Tucker, personal communication). Census Methods In 1986–1987 one of us (W.F.L.) and 2 experienced field assistants used standardized spotlighting methods to repeatedly survey nocturnal mammals in 20 study sites (Fig. 1): 10 forest fragments ranging from 1.4 to 590 ha in area, 7 “controls” in intact but selectively logged rainforest, and 3 corridors along streams (see Laurance 1990, 1991a, 1997 for details). All censuses were conducted along forest edges or old logging tracks. The main species encountered were the red-legged pademelon (Thylogale stigmatica), a scansorial rainforest wallaby, and 5 arboreal folivorous mammals, the coppery brushtail possum (Trichosurus vulpecula johnstoni), lemuroid ringtail (Hemibelideus lemuroides), green ringtail (Pseudochirops archeri) and Herbert River ringtail possums (Pseudochirulus herbertensis), and Lumholtz’s tree-kangaroo (Dendrolagus lumholtzi). Two of us (W.F.L., S.G.L.) returned to the study area in 2006–2007 to resurvey the same sites. We used Conservation Biology Volume 22, No. 5, 2008 1156 Long-Term Dynamics of Mammals Figure 1. Map of study area in north Queensland in 2006–2007. Four control sites in intact forest are A, C, D, and G; 10 forest fragments ranging from 1.4 to 590 ha in area are numbers 1–10; and 4 regrowth “corridors” along streams are a–d. identical methods and surveyed the same routes as in 1986–1987. All 10 fragments were resurveyed, but only 4 of the 7 control sites were accessible because of heavy cyclone damage. In addition, a landowner refused us access to 1 corridor, so we replaced this with 2 other nearby corridors that we repeatedly surveyed by spotlighting in the early 1990s (Laurance & Laurance 1999). To help standardize comparisons between 1986–1987 and 2006– 2007, we included only surveys conducted between the mid-dry season (September–October) and early wet season (January–February). Sampling effort was similar between the 2 intervals. In 1986–1987 each study site was surveyed 4–7 times (104.0 h total), yielding 1035 animal detections, whereas in 2006–2007 each site was surveyed 5 times (124.1 h total), yielding 858 animal detections. Prior work suggests that 4–5 spotlighting surveys are sufficient to obtain stable abundance estimates for most species (Laurance & Laurance 1996). All surveys were conducted on foot between 2000 and 0100 with hand-held spotlights and binoculars to identify animals. Tree-kangaroos and pademelons were also Conservation Biology Volume 22, No. 5, 2008 identified when heard nearby, because the former has a unique escape behavior (plummeting down from trees and then loudly bounding away) and the latter gives distinctive warning thumps when disturbed. The 2 observers alternated among study sites and habitat types to minimize effects of observer bias, with a minimum 10-day interval between successive surveys of the same site. For each study site we determined the abundance of each species (mean number of individuals detected per hour) in 1986–1987 and 2006–2007. For each animal we recorded species, time of observation, estimated height of the animal, estimated horizontal distance of the animal from the forest or road edge, whether the animal was accompanied by conspecifics, and an age-class estimate for the animal, when possible (see Laurance 1990 for details). Landscape Predictors Six quantitative predictors were recorded for each habitat fragment on the basis of 1986 aerial imagery and other data summarized in Laurance (1990), updated with 1157 Laurance et al. recent (August 2006) aerial imagery. These variables were log 10 fragment area; a continuous index of fragment shape; a simple index of topographic diversity (1, flat or nearly flat terrain; 2, gentle or rolling slopes; 3, moderately steep slopes; 4, steep slopes or sharply dissected terrain); a composite index of fragment isolation (integrating 3 intercorrelated measurements of distance to nearby intact forest and large fragments); a “corridorgap code” that described the largest gap in tree-cover along stream-corridors linking each fragment to nearby intact forest (1, no gap in the corridor; 2, 10- to 50m gap; 3, 50- to 100-m gap; 4, 100- to 200-m gap; 5, 200- to 500-m gap; and 6, >500-m gap); and an index of fragment age in 2006 (1, <30 years since isolation; 2, 30–40 years; 3, 40–55 years; 4, >55 years). Detailed descriptions of these predictors are provided in Laurance (1990). Data Analysis We evaluated changes in mammal abundances and species richness in 2 ways. First, we used repeatedmeasures analysis of variance (ANOVA) to test for changes in each response variable between the 2 time intervals (1986–1987 vs. 2006–2007) and among 4 habitat types (4 controls, 5 small [<20-ha] fragments, 5 large [>20-ha] fragments, and 4 corridors). Second, within each habitat type, we used paired t tests to contrast each variable between the 1986–1987 and 2006–2007 periods. Most response variables were log(x+1) transformed to improve data normality and reduce heteroscedasticity, with Wilk-Shapiro tests used to assess data normality. We used best subsets and multiple regressions to assess the influence of landscape predictors on mammal species richness. None of the predictors was intercorrelated strongly enough (R2 < 50%) to produce significant colinearity effects in the multiple-regression models. Performance of the final regression model was assessed by comparing the standardized residuals to fitted values and to each significant predictor. Raw data on foraging heights of the 5 arboreal species could not be located for 1986–1987, so comparisons among habitat types and time periods were conducted by comparing their means and 95% CIs (x̄ ± 1.96 SE), with the 1986–1987 data generated from descriptive statistics in Laurance (1990). Although such comparisons assume a normal data distribution, foraging heights of the arboreal species in 2006–2007 exhibited only modest departures from normality (Wilk-Shapiro statistic = 0.88–0.96), suggesting this assumption is not unreasonable. Raw data on animal detectability (horizontal distance of each individual from the forest edge or logging track) also were lost for 1986–1987, but key inferences were possible with summary statistics in Laurance (1990). Results Animal Detectability In 1986–1987, 93.7% of all animals were detected within 15 m of forest or logging-track edges. Mean and maximum detection distances were 5.7 and 40 m, respectively. Detectability increased slightly in 2006–2007. Of all animals detected, 85.5% were within 15 m of edges, and the mean and maximum detection distances were 7.2 and 70 m, respectively. Thus, approximately 8% more individuals were detected in 2006–2007 relative to 1986– 1987, mostly >15 m from transects. We did not attempt to correct for this difference (see Discussion). Animal detectability did not differ between intact forest versus forest fragments and corridors (F 1,828 = 0.01, p = 0.91), despite large differences in mean detection distances among species (F 5,828 = 7.65, p < 0.001; 2-way ANOVA with log-transformed distance data for 2006– 2007). In pairwise comparisons following a one-way ANOVA (F 5,834 = 18.81, p < 0.0001), tree-kangaroos, pademelons, and lemuroid ringtails were detected at larger distances than coppery brushtails, green ringtails, and Herbert River ringtails (p ≤ 0.002). Tree-kangaroos and pademelons were often identified by sound, which increased their detection distances, whereas lemuroid ringtails have an especially bright eyeshine. In addition, Herbert River ringtails had a larger detection distance than green ringtails (p = 0.04) ( Tukey’s tests), which have a dim eyeshine and cryptic coloration. Species Abundances Repeated measures ANOVAs revealed strong effects of both habitat type and time on mammal community composition (Table 1). Lemuroid ringtail possums, Lumholtz’s tree-kangaroos, and red-legged pademelons varied significantly in abundance among the 4 habitat types (controls, large fragments, small fragments, and corridors), as did overall mammal species richness. Effects of time were significant for coppery brushtail possums, green ringtail possums, tree-kangaroos, and pademelons, and Herbert River ringtails exhibited a significant time-habitat interaction. Results of paired t tests comparing mammal assemblages in each habitat type and time (Table 2) showed that, relative to 1986–1987, tree-kangaroo abundance declined significantly in intact forest in 2006–2007. Herbert River ringtails declined in abundance in large fragments, whereas coppery brushtails, green ringtails, and red-legged pademelons all declined in small fragments. In addition, tree-kangaroos and pademelons exhibited marginally nonsignificant (p ≤ 0.08) declines in large fragments. No species changed in abundance in corridors, although 2 species (lemuroid ringtails and pademelons) were never detected in corridors. A tendency for many species to decline in abundance over time, Conservation Biology Volume 22, No. 5, 2008 1158 Long-Term Dynamics of Mammals Table 1. A comparison with repeated-measures analyses of variance of mammal abundances and species richness between 2 time intervals (1986–1987 and 2006–2007) and among 4 habitat categoriesa . Time Response variable Coppery brushtail possum Green ringtail possumb Herbert River ringtail possumb Lemuroid ringtail possumb Lumholtz’s tree-kangaroob Red-legged pademelonb Species richness F 1,14 p 7.87 9.41 0.18 0.95 11.32 15.63 3.94 Time∗ habitat interaction Habitat F 3,14 0.014 0.008 0.68 0.35 0.005 0.001 0.067 0.25 0.21 1.07 144.51 4.12 5.36 13.59 p F 3,14 p 0.86 0.89 0.39 < 0.001 0.027 0.011 < 0.001 0.32 0.36 5.53 2.73 1.90 2.21 1.35 0.81 0.78 0.010 0.084 0.18 0.13 0.30 a Intact b Data forest; small (<20-ha) forest fragments; large (>20-ha) fragments; riparian corridors. log(x+1) transformed. in 2006–2007. Regressions of species richness against fragment area were significant, or nearly so, for both 1986–1987 (F 1,8 = 4.21, R2 = 34.5%, p = 0.074) and 2006–2007 (F 1,8 = 7.01, R2 = 46.7%, p = 0.029), but slopes of the species–area relationship became much steeper over time (Fig. 3; linear regressions with logtransformed axes). Unlike forest fragments, species richness was constant over time in intact forest (5.8 [0.5] species) and varied little over time in corridors (2.5 [1.0] vs. 2.5 [0.6] species in 1986–1987 and 2006–2007, respectively). On the basis of 1986–1987 data, 3 of the 6 landscape variables were selected as predictors of mammal species richness in fragments, with best-subsets and multiple regressions. Log species richness increased with log fragment area and declined with larger corridor gaps and greater fragment isolation (F 3,9 = 18.08, R2 = 90.0%, p = 0.002). These same 6 landscape variables were used to predict species richness in 2006–2007, but the corridor-gap variable was modified for 4 fragments (numbers 3, 5, 7, and 8; Fig. 1) to reflect changes in the surrounding matrix vegetation. As before, log species especially in forest fragments, was apparent in the comparison of species distributions between 1986–1987 and 2006–2007 (Fig. 2). The most vulnerable species we studied was the lemuroid ringtail possum (Fig. 2). In 1986–1987 it was detected in only a single, 27-ha forest fragment (fragment 7, Fig. 1), and its overall abundance in fragments and corridors declined by 98.2% relative to intact forest. In 2006–2007 it was never observed in the 27-ha fragment, but a single individual was detected in a 12.8-ha fragment (fragment 5, Fig. 1) located just 220 m from intact forest. Compared to intact forest, its abundance in fragments and corridors decreased by 99.98%. Species Richness Mammal species richness declined significantly over time in forest fragments. Effects were not significant when small (p = 0.099) and large (p = 0.18) fragments were evaluated separately (Table 2), but were significant when data from all fragments were pooled (t = 2.86, df = 9, p = 0.019; paired t test). Forest fragments averaged 4.4 species (SD 1.0) in 1986–1987, but just 3.5 (1.4) species Table 2. Paired t tests contrasting species abundances and species richness of rainforest mammals in 4 habitat categoriesa between 1986–1987 and 2006–2007b . Controlsc Response variable Coppery brushtail possum Green ringtail possume Herbert River ringtail possume Lemuroid ringtail possume Lumholtz’s tree-kangarooe Red-legged pademelone Species richness a Intact t −2.88 −1.60 0.61 1.64 −8.70 −2.11 0.00 Large fragmentsd p 0.064 0.21 0.58 0.20 0.003 0.13 1.00 t −0.69 −0.88 −4.73 −1.00 −2.34 −2.59 −1.63 p 0.53 0.43 0.009 0.37 0.080 0.061 0.18 forest; small (<20-ha) forest fragments; large (>20-ha) fragments; riparian corridors. negative t statistic indicates the parameter declined over time. c n = 4 sites. d n = 5 sites. e Data log(x+1) transformed. bA Conservation Biology Volume 22, No. 5, 2008 Small fragmentsd t −2.77 −2.84 0.79 1.00 0.32 −3.11 −2.14 p 0.050 0.047 0.48 0.37 0.77 0.036 0.099 Corridorsc t −1.02 −1.63 2.25 — −1.22 — 0.00 p 0.38 0.20 0.11 — 0.31 — 1.00 1159 Laurance et al. Figure 2. Contrasting abundances (mean no. individuals/h) of 6 mammal species across 18 study sites in tropical Queensland, between 1986–1987 and 2006–2007. The diagonal lines show y = x. richness rose with log fragment area, declined with greater corridor gaps and fragment isolation, and declined with topographic variability (F 3,9 = 13.45, R2 = 91.5%, p = 0.007). In these models, fragment area accounted for less than half (35–47%) of the total variability in species richness, with measures of fragment connectivity (corridor gaps, fragment isolation) accounting for another 27–56% of the total variability. Foraging Heights In 1986–1987 significant differences in mean foraging heights were evident among the 5 arboreal species (Fig. 4). In intact forest, lemuroid ringtails and green ringtails foraged at greater heights than did coppery brushtails and Herbert River ringtails, which in turn foraged at greater heights than the larger-bodied tree-kangaroos. In addition, foraging heights of coppery brushtails, green ringtails, and Herbert River ringtails all shifted downward in fragments and corridors, relative to intact forest (Laurance 1990). Despite modest differences between 1986–1987 and 2006–2007, overall patterns were similar between the 2 intervals. No species significantly altered their foraging heights over time in either fragmented or intact forest, on the basis of comparisons of 95% CIs (Fig. 4). In 2006–2007 foraging heights varied among arboreal species in intact forest (F 4,449 = 15.57, p < 0.0001; one-way ANOVA), with lemuroid ringtails foraging higher than all species except green ringtails (p < 0.05, Tukey’s test). In addition, coppery brushtails foraged at lower heights in fragments and corridors than in intact forest (t = 2.21, df = 230, p = 0.028), but differences for the other species were not significant (all 2-sample t tests). Discussion Animal Detectability The spotlighting surveys we conducted in 1986–1987 and 2006–2007 were both preceded by major cyclones that caused considerable forest disturbance in our study area Conservation Biology Volume 22, No. 5, 2008 1160 Figure 3. Species–area relationships for mammals in rainforest fragments in tropical Queensland; contrasting patterns in 1986–1987 (species = 3.40 area0.08 ) and 2006–2007 (species = 1.67area0.22 ). (Laurance 1991b; Laurance & Curran 2008; Turton 2008). Did these storms affect animal detectability? It has been suggested that, at least for arboreal possums and treekangaroos in intact forest, the principal effect of the 2006 cyclone was to increase animal detectability because the damaged forest was more open than undisturbed forest (Kanowski et al. 2008). The 2006 cyclone caused greater forest damage than the one in 1986, and we estimate this damage enhanced animal detectability by approximately 8% in 2006–2007 relative to 1986–1987. Correcting for this difference had minimal effects on our statistical analyses, and we elected to use uncorrected data because of a general concern about potentially creating artificial structure in our data set. One caveat is that the estimated abundance declines in several species (Table 2) may be conservative, because animal detectability was somewhat higher in 2006–2007 than in 1986–1987. Dynamics of Mammal Assemblages During our initial survey in 1986–1987, most forest fragments in our study area were already 20–50 years old. Long-Term Dynamics of Mammals Over the following 20 years, we documented 5 general changes in mammal assemblages. First, species richness of mammals in fragments, which was already depressed relative to intact forest, declined even further over time (Fig. 3). This decline was greater in small than large fragments (Table 2), resulting in a steeper slope of the species–area relationship in 2006– 2007 (z = 0.22) than in 1986–1987 (z = 0.08). This erosion of species richness mainly resulted from an absence of green ringtail possums and Lumholtz’s tree-kangaroos in several smaller (<20-ha) fragments during our latter survey (Fig. 2). Hence, even in a fragmented landscape that had experienced almost no additional deforestation, species richness of larger mammals continued to decline over time. This decline is consistent with most models of extinction kinetics, which predict a relatively rapid loss of forest-dependent species in recently isolated fragments (Terborgh 1975; Brooks et al. 1999; Watson 2002). Eventually, the pace of species loss is expected to decrease as fragments approach an “equilibrium” species richness. In simple island-biogeographic models (MacArthur & Wilson 1967), this equilibrium is determined solely by the size and isolation of the fragment, which are assumed to govern the rates of species colonization and extinction in fragments. In reality, however, species losses and gains in fragments can be strongly influenced by additional factors such as edge effects (Laurance et al. 2002); human activities, such as logging and hunting (Laurance & Cochrane 2001); and the dynamics of the surrounding matrix vegetation (Gascon et al. 1999; Laurance & Laurance 1999). Second, 4 mammal species, the coppery brushtail possum, green ringtail possum, Herbert River ringtail possum, and red-legged pademelon, declined significantly in abundance in small (<20-ha) or large (>20-ha) forest fragments, and the Lumholtz’s tree-kangaroo exhibited a nearly significant decline (p = 0.08) in large fragments (Table 2). Among these species, the tree-kangaroo and Herbert River ringtail are considered the most vulnerable to fragmentation (Pahl et al. 1988; Laurance 1990), the former because of its low population density, vulnerability to roadkill, and predation by domestic dogs and dingos Figure 4. Mean foraging heights of arboreal mammals in intact forest (controls) and in fragments and stream corridors (fragments) in 1986–1987 (dark-gray columns) and 2006–2007 (light-gray columns). Error bars show 95% CIs. Conservation Biology Volume 22, No. 5, 2008 1161 Laurance et al. (Newell 1999), and the latter because it is almost strictly arboreal and highly reticent to cross open habitat (Laurance 1990). In 1986–1987 coppery brushtails and green ringtails were at least as abundant in fragments as in intact forest, evidently as a result of their ability to cross relatively small expanses of open ground and dietary requirements that include a mix of secondary- and primary-forest plant species (Procter-Gray 1984; Laurance 1990; Jones et al. 2006). Similarly, pademelons are often abundant in forest fragments and are considered an edge-favoring species (Vernes et al. 1995). It is unknown why abundances of these species declined in 2006–2007, although many fragments were heavily disturbed by the 2006 cyclone (Laurance & Curran 2008). We suggest that this cyclone had a negative effect on mammal communities, at least in forest fragments where damage was intense and where resident mammal populations are small, despite a lack of evidence for such an effect in intact forest (Kanowski et al. 2008). Third, Lumholtz’s tree-kangaroos declined significantly in our 4 intact-forest sites (Table 2). The most likely explanation, we believe, is that the spotlighting routes we used in intact forest had previously been selectively logged, promoting a temporary flush of pioneer plant species, some of which are favored by tree-kangaroos (ProcterGray 1984; Newell 1999). Except for control D (which partly encompassed private land), logging operations in our intact-forest sites were halted in 1988 when the region was protected as a World Heritage area, and young pioneer plants declined in these areas. It is also not inconceivable that sampling variation contributed to the apparent decline of tree-kangaroos, given their relative rarity, but this seems a less likely cause than successional changes in intact forest. Fourth, the lemuroid ringtail possum has nearly vanished from fragmented and regrowth forest in our study area, with its abundance being just a tiny fraction (0.02%) of that in nearby intact forest. The lemuroid ringtail exhibits a suite of traits, such as being strictly arboreal, feeding almost entirely on the leaves of primaryforest trees, and requiring a hollow tree-cavity for daytime denning that makes it particularly vulnerable to forest fragmentation (Pahl et al. 1988; Laurance 1990). A key factor that predisposes lemuroids to local extinction is their strong reliance on primary forest, which means that populations in fragments are entirely isolated and therefore highly vulnerable to random demographic and genetic effects (Laurance 1990, 1991a). In a survey of 36 potential faunal corridors on the Atherton Tableland, the lemuroid ringtail was only ever detected in wide (100–300 m), primary-forest corridors that were directly linked to nearby intact forest (Laurance & Laurance 1999). Finally, our most surprising finding was the detection of a lone lemuroid ringtail in a small (12.8-ha) fragment located just 220 m from intact forest (fragment 5, Fig. 1). It seems inconceivable that a relict population of lemuroids persisted in this small fragment, which was isolated since at least 1951 (Pahl 1979). With its brilliant eyeshine, the lemuroid is the most easily detected of all the mammal species we encountered, yet it was never previously detected in the fragment despite repeated spotlighting surveys in 1979 (Pahl et al. 1988), 1986–1987 (Laurance 1990) and 1991–1992 (Vernes 1994). In 1986 only a narrow (generally <20 m wide), discontinuous band of regrowth linked the fragment with nearby intact forest. By 2006, however, this band had coalesced into a much wider (100–200 m wide) regrowth mosaic with a core of tall (>25 m) secondary-forest trees. Although lemuroids have never been detected in regrowth forest (Laurance 1990, 1991a; Laurance & Laurance 1999), the most plausible explanation, we believe, is that one or perhaps a few individuals recolonized the fragment from nearby intact forest. Heavy cyclone damage might have contributed to this by prompting some unusual animal-dispersal movements (J. W. Winter, personal communication). Conservation Implications Our results have 2 implications of general importance. The first is that fragment connectivity appears to play a key role in the maintenance of mammal species richness. In multiple-regression models, 2 landscape variables describing the size of discontinuities in stream corridors and the distance of fragments from other forest areas explained from 27% (1986–1987) to 56% (2006– 2007) of the total variation in species richness. Moreover, among these mammal species there is a strong association between matrix tolerance and survival in fragmented forests, with corridor-using species persisting in many fragments and corridor-avoiding species tending to decline or disappear (Laurance 1990, 1991a, 1994, 1997). These trends highlight the importance of faunal corridors for partially mitigating the effects of habitat fragmentation, a conclusion that accords with other studies of arboreal-mammal assemblages in Australia (Lindemayer et al. 1994; Lindenmayer & Possingham 1996; Downes et al. 1997). We believe corridor effectiveness is likely to interact strongly with fragment isolation and that the most strongly forest-dependent species, such as the lemuroid ringtail possum, are more likely to use corridors to traverse short rather than long distances (Laurance & Laurance 2003). For vulnerable mammal species in tropical Queensland, the best corridors will be wide (>200 m) and continuous; composed of primary forest (or at least mature, species-rich secondary forest); and occur at higher (>750 m) elevations (Laurance & Laurance 1999). Targeted reforestation efforts are increasingly being used to accelerate the establishment of faunal corridors in fragmented landscapes (Goosem & Tucker 1995; Tucker & Murphy 1997; Lamb et al. 2005). Conservation Biology Volume 22, No. 5, 2008 1162 Long-Term Dynamics of Mammals it suggests that large, intact forest tracts are required to ensure the long-term persistence of lemuroid ringtails and other vulnerable wildlife. Moreover, such models do not consider the strong likelihood that global warming and atmospheric changes will diminish habitat quality for cool-adapted, upland-endemic species, such as the lemuroid ringtail (Hilbert et al. 2001, 2004; Kanowski 2001; Williams et al. 2003), in which case minimum critical areas for population survival might well be much larger. Acknowledgments Figure 5. Estimated time to local extinction for the lemuroid ringtail possum (Hemibelideus lemuroides) in tropical Queensland as a function of fragment area. Error bars indicate the possible range of fragment ages. The second key implication is that the most vulnerable species in this region, such as the lemuroid ringtail possum, musky rat-kangaroo (Hypsiprimnodon moshcatus), spotted-tailed quoll (Dasyurus maculatus), Atherton antechinus (Antechinus godmani), and Southern Cassowary (Casuarius casuarius), have disappeared or declined in forest fragments (<600 ha in area) with surprising rapidity (Laurance 1997). The kinetics of local extinction are especially well documented for lemuroid ringtails because of their high detectability; the wellknown history of forest fragments in our study area (see Pahl 1979; Laurance 1990, and references therein); and because arboreal mammals in our fragments were repeatedly surveyed by spotlighting in 1979 (Pahl et al. 1988), 1986–1987 (Laurance 1990, 1991a), and 2006–2007 (this study). These observations reveal that lemuroids disappeared from a small (1.4-ha) fragment in 3–9 years (Laurance 1990), from a medium-sized (27-ha) fragment in 10–29 years (this study), and from 2 larger (43–75 ha) fragments in 35–61 years (Laurance 1990). This pattern suggests a strong effect of fragment area on time to extinction (Fig. 5), which is in agreement with other studies (Brooks et al. 1999; Ferraz et al. 2003). The curvilinear relationship in Fig. 5 is best fitted by a power function (R2 = 84%); linear, logarithmic, and exponential models provided weaker (R2 < 81%) fits. Extrapolating from this curve (Fig. 5) implies that an isolated forest fragment of approximately 300 ha would be required to increase persistence time of lemuroids to a century, whereas a far larger fragment (approximately 24,000 ha) is needed to increase persistence time to a millennium. Obviously this is very rough reckoning, but Conservation Biology Volume 22, No. 5, 2008 G. Ferraz, M. Goosem, J. Kanowski, C. Sekercioglu, N. Sodhi, K. Vernes, and an anonymous referee commented on the manuscript. We thank the National Geographic Society and Smithsonian Tropical Research Institute for financial support; the Wet Tropics Management Authority and Queensland Department of Natural Resources and Water for access to aerial imagery; and CSIRO Tropical Forest Research Centre and N. Tucker, T. Tucker, R. Ewers, J. Kanowski, S. Goosem, B. Petit, and M. Stott for logistical support. Literature Cited Brook, B. W., N. S. Sodhi, and P. K. L. Ng. 2003. Catastrophic extinctions follow deforestation in Singapore. Nature 424:420– 423. Brooks, T. M., S. L. Pimm, and J. O. Oyugi. 1999. Time lag between deforestation and bird extinction in tropical forest fragments. Conservation Biology 13:1140–1150. Corlett, R. T., and I. M. Turner. 1997. Long-term survival in tropical forest remnants in Singapore and Hong Kong. Pages 333–345 in W. F. Laurance and R. O. Bierregaard, editors. Tropical forest remnants: ecology, management and conservation of fragmented communities. The University of Chicago Press, Chicago, Illinois. Curran, L. M., S. Trigg, A. McDonald, D. Astiani, Y. Hardiono, P. Siregar, I. Caniago, and E. Kasischke. 2004. Lowland forest loss in protected areas of Indonesian Borneo. Science 303:1000–1003. DeFries, R., A. Hansen, A. C. Newton, and M. C. Hansen. 2005. Increasing isolation of protected areas in tropical forests over the past twenty years. Ecological Applications 15:19–26. Downes, S. J., K. A. Handasyde, and M. A. Elgar. 1997. The use of corridors by mammals in fragmented Australian eucalypt forests. Conservation Biology 11:718–726. Feeley, K. J., and J. W. Terborgh. 2006. Habitat fragmentation and effects of herbivore (howler monkey) abundances on bird species richness. Ecology 87:144–150. Ferraz, G., G. J. Russell, P. C. Stouffer, R. O. Bierregaard, S. L. Pimm, and T. E. Lovejoy. 2003. Rates of species loss from Amazonian forest fragments. Proceedings of the National Academy of Sciences of the United States of America 100:14069–14073. Gascon, C., T. E. Lovejoy, R. O. Bierregaard, J. R. Malcolm, P. C. Stouffer, H. Vasconcelos, W. F. Laurance, B. Zimmerman, M. Tocher, and S. Borges. 1999. Matrix habitat and species persistence in tropical forest remnants. Biological Conservation 91:223–229. Goosem, S. P., and N. I. J. Tucker. 1995. Repairing the rainforest: theory and practice of rainforest reestablishment in north Queensland’s Wet Tropics. Wet Tropics Management Authority, Cairns, Australia. 1163 Laurance et al. Hilbert, D. W., M. Bradford, T. Parker, and D. A. Westcott. 2004. Golden bowerbird (Prionodura newtoniana) habitat in past, present and future climates: predicted extinction of a vertebrate in tropical highlands due to global warming. Biological Conservation 116:367– 377. Hilbert, D. W., B. Ostendorf, and M. S. Hopkins. 2001. Sensitivity of tropical forests to climate change in the humid tropics of north Queensland. Austral Ecology 26:590–603. Jones, K. M. W., S. J. MacLagan, and A. D. Krockenberger. 2006. Diet selection in the green ringtail possum (Pseudochirops archeri): a specialist folivore in a diverse forest. Austral Ecology 31:799– 807. Kanowski, J. 2001. Effects of elevated CO 2 on the foliar chemistry of seedlings of two rainforest trees from north-east Australia: implications for folivorous marsupials. Austral Ecology 26:165–172. Kanowski, J., M. S. Hopkins, H. Marsh, and J. W. Winter. 2001. Ecological correlates of folivore abundance in north Queensland rainforests. Wildlife Research 28:1–8. Kanowski, J., J. W. Winter, and C. P. Catterall. 2008. Impacts of Cyclone Larry on arboreal folivorous marsupials endemic to upland rainforests of the Atherton Tableland, Australia. Austral Ecology 33:541–548. Lamb, D., P. D. Erskine, and J. A. Parotta. 2005. Restoration of degraded tropical forest landscapes. Science 310:1628–1632. Laurance, S. G., and W. F. Laurance. 1999. Tropical wildlife corridors: use of linear rainforest remnants by arboreal mammals. Biological Conservation 91:231–239. Laurance, S. G. W., and W. F. Laurance. 2003. Bandages for wounded landscapes: faunal corridors and their roles in wildlife conservation in the Americas. Pages 313–325 in G. Bradshaw and P. Marquet, editors. How landscapes change: human disturbance and ecosystem fragmentation in the Americas. Springer-Verlag, New York. Laurance, W. F. 1990. Comparative responses of five arboreal marsupials to tropical forest fragmentation. Journal of Mammalogy 71:641– 653. Laurance, W. F. 1991a. Ecological correlates of extinction proneness in Australian tropical rainforest mammals. Conservation Biology 5:79– 89. Laurance, W. F. 1991b. Edge effects in tropical forest fragments: application of a model for the design of nature reserves. Biological Conservation 57:205–219. Laurance, W. F. 1994. Rainforest fragmentation and the structure of small mammal communities in tropical Queensland. Biological Conservation 69:23–32. Laurance, W. F. 1997. Responses of mammals to rainforest fragmentation in tropical Queensland: a review and synthesis. Wildlife Research 24:603–612. Laurance, W. F. 2007. Have we overstated the tropical biodiversity crisis? Trends in Ecology & Evolution 22:65–70. Laurance, W. F., and R. O. Bierregaard Jr., editors. 1997. Tropical forest remnants: ecology, management and conservation of fragmented communities. The University of Chicago Press, Chicago, Illinois. Laurance, W. F., and M. A. Cochrane. 2001. Synergistic effects in fragmented landscapes. Conservation Biology 15:1488–1489. Laurance, W. F., and T. J. Curran. 2008. Impacts of wind disturbance on fragmented tropical forests: a review and synthesis. Austral Ecology 33:399–408. Laurance, W. F., and S. G. W. Laurance. 1996. Responses of five arboreal marsupials to recent selective logging in tropical Australia. Biotropica 28:310–322. Laurance, W. F., T. E. Lovejoy, H. L. Vasconcelos, E. M. Bruna, R. K. Didham, P. C. Stouffer, C. Gascon, R. O. Bierregaard, S. G. Laurance, and E. Sampiao. 2002. Ecosystem decay of Amazonian forest fragments: a 22-year investigation. Conservation Biology 16:605–618. Lindenmayer, D. B., R. B. Cunningham, C. F. Donnelly, B. E. Triggs, and M. Belvedere. 1994. Factors influencing the occurrence of mammals in retained linear strips (wildlife corridors) and contiguous stands of montane ash forest in the central highlands of Victoria, southeastern Australia. Forest Ecology and Management 67:113–133. Lindenmayer, D. B., and J. Fischer. 2006. Habitat fragmentation and landscape change: an ecological and conservation synthesis. Island Press, Washington, D.C. Lindenmayer, D. B., and H. P. Possingham. 1996. Modelling the relationships between habitat connectivity, corridor design and wildlife conservation within intensively logged wood production forests of south-eastern Australia. Landscape Ecology 11:79–105. Lovejoy, T. E., et al. 1986. Edge and other effects of isolation on Amazon forest fragments. Pages 257–285 in M. E. Soulé, editor. Conservation biology: the science of scarcity and diversity. Sinauer Associates, Sunderland, Massachusetts. MacArthur, R. M., and E. O. Wilson. 1967. The theory of island biogeography. Princeton University Press, Princeton, New Jersey. Mayaux, P., P. Holmgren, F. Achard, H. Eva, H.-J. Stibig, and A. Branthomme. 2005. Tropical forest cover change in the 1990s and options for future monitoring. Philosophical Transactions of the Royal Society of London B 360:373–384. Newell, G. R. 1999. Home range and habitat use by Lumholtz’s treekangaroo (Dendrolagus lumholtzi) within a rainforest fragment in north Queensland. Wildlife Research 26:129–145. Pahl, L. I. 1979. A study of the distributions of selected marsupial species (families: Phalangeridae, Petauridae, and Macropodidae) in a tropical rain forest reduced and fragmented by man. BS honors thesis. James Cook University, Townsville, Australia. Pahl, L. I., J. W. Winter, and G. Heinsohn. 1988. Variation in responses of arboreal marsupials to fragmentation of tropical rainforest in north eastern Australia. Biological Conservation 46:71–82. Procter-Gray, E. 1984. Dietary ecology of the coppery brushtail possum, green ringtail possum and Lumholtz’s tree-kangaroo in north Queensland. Pages 115–128 in A. H. Smith and I. D. Hume, editors. Possums and gliders. Australian Mammal Society, Sydney. Rudel, T. K. 2005. Tropical forests: regional paths of destruction and regeneration in the late twentieth century. Columbia University Press, New York. Stouffer, P. C., R. O. Bierregaard Jr., C. Strong, and T. E. Lovejoy. 2006. Long-term landscape change and bird abundance in Amazonian rainforest fragments. Conservation Biology 4:1212–1223. Terborgh, J. 1975. Faunal equilibria and the design of wildlife preserves. Pages 369–380 in F. Golley and E. Medina, editors. Tropical ecological systems: trends in terrestrial and aquatic research. Springer-Verlag, New York. Terborgh, J., et al. 2001. Ecological meltdown in predator-free forest fragments. Science 294:1923–1926. Tilman, D., J. Fargione, B. Wolff, C. D’Antonio, A. Dobson, R. Howarth, D. Schindler, W. Schlesinger, D. Simberloff, and D. Swackhamer. 2001. Forecasting agriculturally driven global environmental change. Science 292:281–284. Tucker, N. I. J., and T. M. Murphy. 1997. The effects of ecological rehabilitation on vegetation recruitment: some observations from the Wet Tropics of north Queensland. Forest Ecology and Management 99:133–152. Turton, S. M., editor. 2008. Ecological impacts of tropical cyclones on Australian terrestrial ecosystems: insights from Cyclones Larry and Monica. Special issue of Austral Ecology 33:365–584. Vernes, K. 1994. Life on the edge: the ecology of the red-legged pademelon Thylogale stigmatica (Gould) (Marsupialia: Macropodidae) in fragmented rainforest in the north Queensland wet tropics. MS thesis. James Cook University, Townsville, Australia. Vernes, K., H. Marsh, and J. W. Winter. 1995. Home-range characteristics and movement patterns of the red- legged pademelon (Thylogale stigmatica) in a fragmented tropical rainforest. Wildlife Research 22:699–708. Watson, D. M. 2002. A conceptual framework for studying species composition in fragments, islands and other patchy ecosystems. Journal of Biogeography 29:823–834. Conservation Biology Volume 22, No. 5, 2008 1164 Wilcox, B. A. 1978. Supersaturated island faunas: a species-age relationship for lizards on post-Pleistocene land-bridge islands. Science 199:996–998. Williams, S. E., E. Bolitho, and S. Fox. 2003. Climate change in Australian tropical rainforests: an impending environmental catastrophe. Proceedings of the Royal Society of London B 270:1887–1892. Williams, S. E., and R. G. Pearson. 1997. Historical rainforest contractions, localized extinctions and patterns of vertebrate endemism in Conservation Biology Volume 22, No. 5, 2008 Long-Term Dynamics of Mammals the rainforests of Australia’s wet tropics. Proceedings of the Royal Society of London B 264:709–716. Winter, J. W. 1988. Ecological specialisation of mammals in Australian tropical and sub-tropical rainforest: refugial or ecological determinism? Proceedings of the Ecological Society of Australia 15:127– 138. Winter, J. W., F. C. Bell, L. I. Pahl, and R. G. Atherton. 1987. Rainforest clearfelling in northeastern Australia. Proceedings of the Royal Society of Queensland 98:41–57.