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Forest Ecology and Management 255 (2008) 1575–1583 www.elsevier.com/locate/foreco

Effects of exotic conifer plantations on the biodiversity of understory plants, epigeal beetles and birds in Nothofagus dombeyi forests Juan Paritsis *, Marcelo A. Aizen Laboratorio Ecotono-CRUB, Universidad Nacional del Comahue, Pasaje Gutierrez 1125, Bariloche, Pcia. Rı´o Negro, Argentina Received 30 May 2007; received in revised form 16 November 2007; accepted 16 November 2007

Abstract Plantations of exotic conifers are a potential threat to natural ecosystems in the Argentinean Patagonia and a major cause of native forest loss in Chile. We examined species diversity and composition of three different functional groups: understory vascular plants, epigeal beetles, and birds, in paired stands of relatively undisturbed Nothofagus dombeyi forest and coniferous plantations. We also characterized the structure of each stand. Exotic plantations generated significant impacts on biodiversity, diminishing species richness, abundance and diversity, and generating modifications in assemblage composition. Replacement of N. dombeyi forests by conifer plantations led to a homogenization of habitat structure at the stand scale. The largest impact was detected on understory plants, followed by the beetle and bird assemblages, supporting the view that the least vagile assemblages are most affected by habitat replacement. The most relevant modifications caused by the plantations on the structure and composition of the studied assemblages were a reduction in evenness in plants and beetles, an increase of exotic species richness from 16 to 29% in plants and from none to one species in birds, and a loss of rare and specialist species in all three assemblages. Our findings suggest that plantations with more open canopy could favor biodiversity by increasing individual abundance and species richness of all three assemblages. # 2007 Elsevier B.V. All rights reserved. Keywords: Biodiversity; Birds; Beetles; Nothofagus; Plantations; Understory plants

1. Introduction The replacement of native forests by exotic tree plantations can cause important changes in diversity and community composition at local and regional scales (Brockerhoff et al., 2001). However, different functional groups of organisms (e.g., taxonomic assemblages, guilds, and trophic levels) perceive this new habitat (i.e., the plantation) in diverse ways according to features such as individual size, movement abilities and trophic requirements. Thereby, the response to this type of disturbance can vary regarding the particular assemblage analyzed (Rymer, 1981). A significant number of studies have evaluated the effect of exotic tree plantations on particular taxonomic assemblages; e.g., plants (Michelsen et al., 1996; Frank and Finckh, 1997; Raffaele and Schlichter, 2000), insects (Saiz and Salazar, 1981; Sinclair and New, 2004; Corley et al., 2006), and birds (Clout and Gaze, 1984; Carlson, 1986; * Corresponding author. Present address: Department of Geography, University of Colorado, Campus Box 260, Boulder, CO 80309-0260, USA. Tel.: +1 30349 22631; fax: +1 30349 27501. E-mail address: [email protected] (J. Paritsis). 0378-1127/$ – see front matter # 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.foreco.2007.11.015

Pomeroy and Dranzoa, 1998; Zalba, 2001). Nevertheless, none of them has simultaneously analyzed the effect of plantations on different functional groups. Many taxonomic assemblages have certain traits that make them good biodiversity indicators. However, the usefulness of any given group as a biodiversity indicator varies with factors such as habitat and community type (Lindenmayer et al., 2000). Thus, the inclusion of various assemblages with pronounced lifehistory differences may indicate broader effects on the biota than when considering one single group of organisms. In addition, the simultaneous consideration of several taxonomic assemblages may provide cues to specific taxon-traits that can explain why some groups of organisms may be more susceptible than others to different kinds of human-driven habitat disturbances. Here we assess the effects of replacing south Andean Nothofagus forest by conifer plantations on the diversity and composition of three contrasting taxonomic assemblages: plants, beetles, and birds. Each one of these taxonomic groups has also a different ecological function in the ecosystem. Plants are primary producers, beetles are a mixture of decomposers, herbivorous and predaceous species, and birds include pollinators, fruit and seed eaters as well as insectivorous species.

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South American Nothofagus forests, stretching along most of southern Chile and the eastern foothills of the Patagonian Andes in Argentina, constitute an important world reserve of biodiversity due to their high incidence of endemic taxa (Heywood, 1995; Armesto et al., 1996) and the occurrence of unique plant–animal mutualisms (Aizen and Ezcurra, 1998; Amico and Aizen, 2000). The flora is highly endemic including 34% of the woody plant genera and three entire angiosperm families occurring exclusively in these southern temperate forests (Aizen and Ezcurra, 1998). The insect fauna bears relict species of great significance to conservation and a relative high number of ancient taxa with Gondwanic affiliation (McQuillan, 1993). Among the birds, there is a higher level of endemism than in many other continental avifaunas, being 41% of the species and 10% of the genera exclusive to this biome (Vuilleumier, 1985). Presently, diverse anthropogenic perturbations threaten the biodiversity of these peculiar forests; being the replacement of native forest habitat by forestry plantations one of the main threats (Armesto et al., 1998). At present, in the Argentinean Patagonia, impacts of exotic conifer plantations on biodiversity at a regional scale may be considered of limited importance due to the relatively small area already planted. However, increased planting of exotic trees in the steppe is being promoted by governmental agencies and the private sector, and many experts foresee a rapid expansion of the areas planted with exotic species (Schlichter and Laclau, 1998). In Chile, where forestry is currently one of the main economical activities, plantations of exotic pines and Eucalyptus spp. have already replaced large areas of native forests covering ca. two million hectares (Lara and Veblen, 1993; CONAF-CONAMA, 1999). In addition, exotic conifers (e.g., Pseudotsuga menziesii) are also invading relatively intact vegetation in northwestern Patagonia (Richardson and Higgins, 1998). Despite the importance of plantations in these areas, few studies have addressed the effects of exotic conifer plantations on native communities, either in Chile or in Argentina (but see Schlichter and Laclau, 1998; Estades and Temple, 1999; Raffaele and Schlichter, 2000; Corley et al., 2006). Thus, assessment of how the native biota reacts to the environmental changes generated by exotic plantations is a critical research need for policy makers. In this study, we compared species diversity and composition of three different assemblages – understory vascular plants, epigeal beetles and birds – in exotic conifer plantations (Pinaceae) and neighboring Nothofagus dombeyi (Mirb.) Blume forest, in which these plantations are immersed. The simultaneous analysis of different taxonomic/functional groups of organisms with disparate life-history traits and spatial perceptions of their surrounding habitat enabled us to carry out a general assessment of the incipient impact of exotic plantations on biodiversity in this southern biome. Because a decrease in habitat heterogeneity has been invoked as an important contributing cause of biodiversity loss in woody biomes (Freemark and Merriam, 1986; Hansen, 2000), we also characterized the structure of each study site. Our specific objectives were to (i) determine how stand structure of the native forest stands differs from that of exotic plantations; (ii)

establish how the replacement of a native forest stand by a plantation affects the biodiversity of plants, epigeal beetles and birds; and (iii) determine whether these three assemblages are differentially affected by this kind of habitat replacement. 2. Methods 2.1. Study area The study was conducted in Nahuel Huapi National Park (Victoria Island) and adjacent private areas (San Pedro peninsula), in northwestern Patagonia (418S, 718W). Four pairs of study sites (i.e., native forest-plantation) were distributed at altitudes ranging from 770 to 850 m with an average annual rainfall of ca.1800 mm (Barros et al., 1983). The tall evergreen angiosperm, Nothofagus dombeyi, dominates the tree canopy, and the bamboo, Chusquea culeou Desv., and shrubs, Schinus patagonicus (Phil.) I. M. Jonhst and Berberis darwinii Hook, the understory layer. In the Victoria Island and San Pedro peninsula, exotic conifer plantations are typically small (ca. five hectares) and of mixed composition (mainly Pseudotsuga menziesii (Mirb.) Franco (20–50% of all planted trees), Pinus radiata D. Don (0– 80%) and Pinus sylvestris L. (0–60%)), and are immersed in a mostly continuous matrix of native N. dombeyi forest. The plantations used in this study are located in the abovementioned areas; two in Victoria Island and two in San Pedro peninsula. Most of the trees in these plantations were planted as small plants; however, for some trees on the Victoria Island it is impossible to determine the stage in which they were planted (Simberloff et al., 2002). Plantations were all composed of mature trees ranging from 35 to 75 years old and were regularly pruned. At each of these four different sites, we selected a pair of habitat units consisting of a control area (native N. dombeyi forest) and an exotic conifer plantation, which were <1 km apart from each other. The native N. dombeyi forest stands chosen for this study as controls were of variable age, ranging from post-fire relatively homogeneous stands, which were paired with the young plantations, to old-growth, which were paired with the old plantations. Common disturbances in the selected N. dombeyi and plantation stands are browsing by introduced red deer (Cervus elaphus L.; in Victoria Island) and European hare (Lepus capensis L.; in San Pedro peninsula), and non-commercial logging of fallen logs. All four study sites were at an average distance of 11.5 km and the two most-distant sites were 19 km apart from each other. All the sites have a similar precipitation regime because they are not significantly apart from each other along the longitudinal west-to-east precipitation gradient (the maximum longitudinal distance between sites was 8 km). Sampling was conducted from January to March 2001 (Austral summer). 2.2. Sampling of stand structure and assemblages We characterized the stand structure of both native N. dombeyi forest stands and plantations by recording tree density

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and canopy cover, and individual tree height and tree diameter at breast height (dbh; 1.30 m). In each stand, we sampled a total of ca. 80 live and dead standing trees (>2 m tall and >5 cm dbh). We randomly distributed 20 m  6 m plots within the stand and measured the height and dbh of all live trees, and the dbh of all dead standing trees within the plot until reaching a total of 80 trees. Thus, total number of plots per stand varied between 5 and 20, depending on tree density. Canopy cover was estimated at the center of each plot using a densiometer. We determined understory vascular plant species composition and richness during mid summer (February 2001). In each stand, we set up a total of thirty 1-m radius plots equally distributed along three parallel 350-m long transects separated 15 m from each other. In each plot, we visually estimated the percentage cover of each species, bare soil and leaf litter, and measured litter depth at the center of the plot. Epigeal (i.e., ground dwelling) beetles were captured using pitfalls filled with a mix of water and propylene glycol (Ausden, 1996). In each stand, we placed eight traps equally spaced along a 350-m transect, collecting the content of each trap at four times during the sampling period. Traps were active in the field during 10-day periods before collecting. In the laboratory, we separated all the adult individuals of the order Coleoptera from the samples and classified them in Recognizable Taxonomic Units (RTU; Oliver and Beattie, 1996; Pik et al., 1999) using a binocular lens (Leica MZ up to 50). Subsequently, RTUs were classified to the lowest identifiable taxonomic level. For recording birds, we used the fixed radius point count method due to the forest structure and the small spatial extent of the surveyed areas (Ralph et al., 1993). In each stand, we established six points (stations) 75 m apart along a linear transect. The same observer recorded all the birds seen or heard during an 8-min period within 20 m from the central point of each station. Bird surveys were done between ca. 6:15 a.m., at sunrise, and 10:30 a.m. We repeated the surveys six times over the sampling season calculating the mean abundance for each species. 2.3. Data analysis Species richness, abundances, and Shannon–Weaver (Peet, 1975) biodiversity indices were calculated for each assemblage. We used paired t-tests to compare the richness and abundance of each assemblage, and stand structural attributes between native forest stands and conifer plantations. Due to inherent problems regarding traditional biodiversity indices, such as the loss of information associated with limited sampling (James and Rathbun, 1981; Gotelli and Graves, 1996); we also calculated rarefied estimates of species richness. Rarefaction analysis allowed us to compare richness and diversity after differences in abundance among samples had been standardized (Simberloff, 1972; Gotelli and Colwell, 2001). Rarefaction analyses were performed using the EcoSim.7 software (Gotelli and Entsminger, 2006). We calculated rank–abundance curves to obtain a detailed and graphic description of abundance distribution for each assemblage in each habitat type (James and Rathbun, 1981).

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Because rank–abundance curves do not consider the identity of the species (i.e., they do not reveal if the most abundant species in one habitat unit was also the most abundant in another habitat unit), we performed Spearman rank correlations to determine if species–abundance rankings varied in a similar way between native forest stands and plantations. To analyze similarity in species composition between native forest stands and plantations we used similarity percentage (PS; Hansen, 2000). This index computes similarity based on species presence/absence and relative species abundance between two samples according to: P 200 Sj¼i minðX j1 ; X j2 Þ PS ¼ PS j¼i ðX j1 þ X j2 Þ where Xj1 is the relative abundance of species j in sample 1 and Xj2 is its abundance in sample 2. We calculated all the possible pairwise comparisons among the eight sampled habitat units, in total 28 PS. Sixteen of these PS indices corresponded to comparisons between native forests and plantations, six to comparisons between native forests and six to comparisons between plantations. The mean PS of each of the three comparisons was compared with a distribution of randomized mean PS indices. The randomized distributions of mean PS values and their respective 2.5 and 97.5 percentiles (Manly, 1991) were obtained from random subgroups drawn from the 28 original PS (based on subsamples with n = 16 for pairwise comparisons between native forests and plantations, and n = 6 for comparisons between forests and between plantations). Randomization procedures were run with Resampling Stats (Simon, 1992). For all the analyses, excepting rarefaction and rank– abundance curves, we used the mean percent cover or frequency, for plants, and mean number of individuals, for beetles and birds, per habitat unit and per species. For the rarefaction and rank–abundance analyses, we used the number of individuals per species, or frequency of occurrence in the case of plants, summed over plots, sampling periods, and stands of the same type (i.e., plantation vs. native forest). Lumping of same habitat-type data across sites here was justified, because we observed similar differential patterns between the two types of stands for any particular site. 3. Results 3.1. Stand structure characterization Nothofagus dombeyi forest stands exhibited greater heterogeneity in their structure than their paired plantations, with some individual trees >1.5 m dbh and >30 m height. On the other hand, plantations presented a relatively homogeneous structure, without individuals >1 m dbh. Mean tree height, dbh and basal area were similar between both habitat units (Table 1). However, variation in dbh, as measured by the CV, revealed that tree size in N. dombeyi forest stands was twice as variable as in the plantations (P < 0.05). Canopy cover was higher in the plantations than in the native forest stands

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Table 1 Summary of the stand characteristics for both habitat types; native forest stands and plantations

Tree height (m) dbh (cm) dbh (coefficient of variation) Basal area (m2/ha) Canopy cover (%) Live tree density (trees/ha) Proportion of dead standing trees (%) dbh of dead standing trees (cm) Leaf litter cover (%) Leaf litter depth (cm) Bare soil (%)

Native forests

Plantations

17.9 (1.7) 45.3 (6.5) 0.69 (0.1)* 109.8 (17.1) 83.5 (1.2)* 562.7 (89.6) 7.5 (2.1) 30.3 (9.3) 75.3 (5.1) 4.7 (0.8) 1.4 (0.5)*

19.4 (1.3) 38.1 (4.8) 0.35 (0.1)* 97.1 (14.9) 92.0 (0.5)* 874 (163.2) 12.5 (2.4) 19.4 (0.7) 88.5 (3.9) 6.7 (0.7) 0.2 (0.1)*

Data are means and standard errors (between parentheses). Significant differences (t test; n = 4; P < 0.05) are marked with an asterisk.

(P < 0.05), and for all three assemblages was inversely and significantly (P < 0.05) correlated to species richness (r values = 0.48 for plants, 0.61 for beetles, and 0.81 for birds) and abundance (r values = 0.68 for plants, 0.79 for beetles, and 0.89 for birds). The proportion of dead standing trees and their dbh did not exhibit significant differences between native forest stands and plantations; however, there was a tendency for larger dead trees (according to dbh) in the native forest stands. Leaf litter cover and depth was, on average, slightly higher in the plantations than in the native forest stands but differences between environments were not statistically significant. However, percentage of bare soil was significantly higher in the native forest stands than in the plantations (Table 1; P < 0.05). 3.2. Species richness, abundance and diversity Exotic conifer plantations replacing native forests generated a decrease in species richness, abundance and diversity of all three studied species assemblages (Table 2). However, the response to this habitat replacement differed among the three assemblages. Plants were the most affected assemblage, beetles showed an intermediate response and birds were the least affected assemblage by habitat transformation. Understory plants exhibited the greatest difference in mean abundance (% cover) between both types of habitats compared to the other two assemblages. Percentage plant cover in the

native forest stands was one order of magnitude greater than percentage plant cover in the plantations. Understory plants also showed a significant decrease in mean species richness in the plantations (Table 2). However, rarefaction analysis indicated that the low plant species richness recorded in the plantations was not merely a consequence of their low abundance: differences in species richness were still apparent after differences in abundance were accounted for (Fig. 1a). Twenty out of the 42 species recorded in the native forests’ understory were exclusively found in this habitat, whereas in the plantations’ understory we found two species only that were growing exclusively in this habitat and both were non-native (i.e., Pinus ponderosa and Acer pseudoplatanus L.). Nearly 16% of all plant species in the understory of the native forest stands were non-natives, while in the plantation this proportion rose to 29%. We collected a total of 2139 individuals of 69 species and morphospecies of epigeal beetles in both habitat types. Mean species richness in the native forest stands was significantly higher than in the plantations (Table 2). Rarefaction analysis confirmed that this result was not an artifact of the difference in the number of collected individuals (Fig. 1b). Overall, we recorded 1256 bird individuals of 21 species belonging to 12 families in both habitat types. Birds showed clear differences in mean abundance between environments. We recorded twice as many individuals in the native forest stands than in the neighboring plantations (Table 2). The most abundant species in the native forest stands were also recorded in the plantations, but with lower abundances. However, there were two species that were absent in the plantations (Scelorchilus rubecula and Tachycineta leucopyga), although they presented fairly high relative abundances in the native forest stands (i.e., more than 1% of all counts). As with the two other assemblages, rarefaction curves showed that the higher species richness in the native forest stands than in the plantations could not be ascribed to differences in abundance between the two habitat types (Fig. 1c). All of the 20 bird species recorded in the N. dombeyi forest stands were native and eight of them were endemic to the South American Austral forests (representing 40% of all the recorded species). In the plantations, one of the 13 species recorded was nonnative (i.e., silver pheasant; Lophura nycthemera) and only three of them were endemic, representing 23% of all the species.

Table 2 Richness, abundance, and diversity of the three assemblages in the native forest stands and in the plantations considering the four sampled sites Plants

Total species richness Mean species richness Mean abundance Shannon–Weaver

Epigeal beetles

Birds

Native forest

Plantation

Native forest

Plantation

Native forest

Plantation

42 24.5 (2.9)* 39.6% (5.5) a * 0.86

24 12.0 (2.0)* 4.0% (1.6) a * 0.40

61 31.2 (3.1)* 535 (119) 0.88

34 13.0 (2.0)* 212 (121) 0.62

20 14.2 (0.6)* 35.7 (3.4)* 0.78

13 9.2 (0.8)* 16.6 (3.4)* 0.70

Mean values correspond to the average of the four different sites (standard errors are between parentheses). The asterisk indicates significant differences (t-test; n = 4; P < 0.05) between native forest stands and plantations. a For plants, mean abundance is the mean percentage of plant cover per sampling plot.

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Fig. 2. Rank–abundance plots comparing both habitat types ((a) plants, (b) epigeal beetles, and (c) birds). Note the logarithmic scale of the Y-axis.

Fig. 1. Rarefaction curves for each of the three assemblages in the native forest stands and in the plantations ((a) plants, (b) epigeal beetles, and (c) birds). Confidence intervals (95%) are provided for the composite curves of the native forest stands and the plantations. Curves represent the expected species richness for different number of individuals (or frequency in the case of plants).

3.3. Species rank–abundance and Spearman rank correlations The composite rank–abundance curve for the plant assemblage from the native forest stands reflects the existence of many species with intermediate abundances (Fig. 2a). In contrast, the steep slope of the rank–abundance found for the plant community inhabiting the plantations reveals the dominance of one species (i.e., new recruits of Pseudotsuga menziesii), and the reduction of species with intermediate

abundances. Spearman rank correlations showed that plant species in plantations diminished their cover to relatively low values, regardless of their percentage cover in the native forests (Fig. 3a). In other words, the ranking in plant species abundance of the continuous forest was not predictive of that in the plantations. The species rank–abundance curves of the epigeal beetles exhibited similar patterns as the understory plants, with the assemblages from the plantations exhibiting a steeper slope than those from the surrounding native forest (Fig. 2b). However, in contrast with plants, beetle assemblages showed a larger proportion of species with very low relative abundances (i.e., <0.1%) in both habitat types. Contrary to the other two groups, birds did not show substantial differences in the distribution of their relative specific abundances between native forest stands and plantations (Fig. 2c). The only relevant difference was the higher number of species with low abundances in the native forests compared to the plantations. Unlike plants, beetles and birds presented high Spearman rank correlation coefficients when comparing the species– abundance ranking of their shared species between plantations and native forest stands (Fig. 3b and c). This indicates that rank–abundance relationships between specific beetles and bird

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Fig. 4. Mean percentage similarity indices for the three types of comparisons: native forest stands vs. plantations (NF vs. Pl), native forest stands among themselves (NF vs. NF) and plantations among themselves (Pl vs. Pl). (a) Plants, (b) epigeal beetles, and (c) birds. Error bars indicate the range of values between the 2.5 and 97.5% percentiles obtained through randomization.

species did not vary regardless of habitat type. Thus, the most abundant beetle or bird species in the native forests were also the most abundant in the plantations. Nevertheless, Spearman correlation coefficients were lower when we included nonshared species between plantations and native forest stands, due to a relatively high number of species with very low abundance in the native forests that were absent in the plantations.

because it was included within the randomly generated 95% confidence interval, we cannot conclude that there are significant differences in plant species composition between the two habitat types (Fig. 4a). Comparisons between habitats in beetle assemblages did not show departures from random expectations either (Fig. 4b). On the other hand, the observed average PS between bird assemblages from the plantations versus the forest stands was significantly lower than randomly expected (Fig. 4c). In general, species assemblages in the native forest were highly similar among the different stands we sampled, and at least for two of the assemblages (plants and birds) more similar than expected by chance (Fig. 4a and c). Observed average similarity indices between plantations did not depart form random expectations for any of the three types of assemblages studied (Fig. 4). Birds presented the highest PS values compared with the other two assemblages, followed by the epigeal beetle and plant assemblages (Fig. 4). PS estimates for plants were generally low for the three types of comparisons.

3.4. Percent similarity in species composition

4. Discussion

The mean percent similarity (PS) between assemblages in the native forest stands versus plantations was relatively low, when contrasted to the two types of intra-habitat comparisons (but epigeal beetles). The PS for plant assemblages between native forest and plantation stands was as low as 4%, but

Substitution of native Nothofagus dombeyi forests by exotic conifer plantations had a significant effect on the study assemblages, diminishing their species richness, abundance, diversity and evenness. However, each group responded in a slightly different fashion to this habitat replacement. In general,

Fig. 3. Spearman correlations between species abundances (percent cover in the case of plants) in the native forests and in the plantations ((a) plants, (b) epigeal beetles, and (c) birds). Two different types of correlations were performed: type 1 considers all the recorded species in both habitat types, and type 2 only the species shared between both habitat types.

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plants were the most affected taxonomic group, followed by epigeal beetles and lastly birds. In addition, understory plant and bird assemblages showed a tendency towards an increase in non-native species in the plantations. Rarefaction and rank– abundance analyses suggest that reductions in species richness caused by plantations relate also to reductions in the evenness of the assemblages, particularly in the case of plants and beetles. Many of these responses have also been documented in different types of substitutions of natural habitats by exotic plantations around the world (Driscoll, 1977; Clout and Gaze, 1984; Michelsen et al., 1996; Frank and Finckh, 1997), but were seldom observed for different taxonomic assemblages studied simultaneously. 4.1. Habitat structure Comparison of the structure of the two habitat types suggests that habitat heterogeneity was higher in the native forest stands than in the plantations. Several authors consider environmental heterogeneity an important factor that promotes biodiversity (Rosenzweig and Abramsky, 1993). Therefore, the higher structural heterogeneity of the native forest stands could be contributing to the higher biodiversity observed in these forests. On the contrary, the structural homogeneity of the plantations might indicate a lower number of available niches. Canopy cover was significantly associated with biodiversity, in both habitat types. Both, native forest stands and plantations with higher percentages of canopy cover presented lower abundances and species richness than those with more open canopies. For exotic plantations in the Patagonian steppe, Corley et al. (2006) also found that more open plantation designs are accompanied by higher ant abundances and species richness. This suggests that canopy cover is a key structural factor when designing and managing plantations to increase biodiversity. 4.2. Habitat replacement and species assemblages In all three assemblages, decreases in species richness in the plantations could not be explained by intrinsic differences in species abundance or sampling effort. If a similar percentage cover (i.e., abundance) were sampled in the case of plants, or individuals in the case of beetles or birds, in both the native forest stands and in the plantations, differences in species richness would still hold. Thus, differences in species diversity between these two types of habitat were not only consistent across the three assemblages studied, but also they seem to be a consequence of habitat conversion. Spearman rank correlations revealed that understory plant species in the plantations suffered a severe reduction in their abundances. This analysis showed that the reduction was unrelated to their abundance in the native forest stands. This plant response suggests that habitat modifications generated by the plantations may diminish the richness and diversity of the understory flora not only through a proportional reduction in species abundances, but also through more selective processes affecting certain species more than others. In addition, none of

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the few native plants in the plantation understory was bearing reproductive structures or any remain of them (J. Paritsis, personal observation). Furthermore, most plants in the plantations were on their first growth stages despite that plantations were 35-year old or older. This suggests that the depauperated plant assemblages found in the plantations depend entirely on the recolonization from the surrounding native forest matrix. It is likely that this recolonization was possible only because of the relatively small size of the plantations. Our findings contrast with studies carried out in the tropics that found a rich native plant understory within pine plantations (Fimbel and Fimbel, 1996; Keenan et al., 1997). It is possible that the higher proportion of understory plants adapted to low-light environments in the tropics compared to these temperate Nothofagus forests is responsible for the differences in plantation’s understory richness between environments. Results from Spearman rank correlations and assemblage similarity analysis indicate that the epigeal beetle community was not selectively affected. In other words, plantations seemed to be disturbing all beetle species with a similar intensity. Therefore, those species with high relative abundances in the native forest stands remained the most abundant in the plantations, but with lower absolute abundances. On the other hand, the least abundant beetle species in the native forest stands completely disappeared from the plantations. Leaf litter structure and composition is regarded as a key factor influencing epigeal beetle biodiversity (Niemela¨ et al., 1992; Koivula et al., 1999). Higher heterogeneity in the leaf litter can favor terrestrial invertebrate diversity (Hansen, 2000; Magura et al., 2005). In addition, the presence of broadleaf leaves in the leaf litter may also promote the diversity of epigeal beetles and other terrestrial invertebrates (Koivula et al., 1999; Hansen, 2000), in part, because broadleaf leaves decompose faster and maintain higher humidity levels than needle leaves (Donoso Zegers, 1995). Leaf litter in the native N. dombeyi forest stands was almost entirely composed by broad leaves and was originated from a wider range of species than the needle leaf litter in the plantations. Thus, the higher epigeal beetle diversity found in the native forest stands can be associated with differences in leaf litter composition between the native forests and the plantations. Birds showed the smallest decrease in species richness in the plantations compared with the other two assemblages. Spearman correlations showed that the bird assemblages found in the plantations are a subgroup of those recorded in the native forest stands in which the rarest species are absent. The tapaculos (Fam: Rhinocryptidae) and the woodpeckers (Fam: Picidae) were the most affected by the replacement of native forest stands by conifer plantations, most likely because of their specific habitat needs. Other studies in different regions of the world also documented that specialist bird species diminished their abundances or were not found in exotic plantations (Clout and Gaze, 1984; Carlson, 1986; Pomeroy and Dranzoa, 1998). Thus, our results provide another piece of evidence to a global phenomenon of the lost of specialist birds due to habitat conversion.

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Despite the small size of the plantations surveyed in this study (ca. 5 ha), changes in biodiversity compared to similar areas in the native forest matrix were dramatic. The plantations were unsuitable or rejected by a large number of species despite they were small fragments surrounded by native forest. In a study on the effects of pine plantations on the avian biodiversity in Chile, less striking changes in bird abundance and richness between native forests fragments and a matrix of plantations were found (Estades and Temple, 1999). It was suggested by the authors that birds tend to adapt to the more common habitat type in the area. Nevertheless, in our case we are not able to predict if larger plantations would sustain higher biodiversity or if, on the contrary, would be more depauperate compared to native habitats. Yet it is clear that many native species are not able to survive in the homogeneous, low-light environment present in the plantations, and at least in the case of most plants, they may not reproduce even if they establish. 4.3. Management implications Unlike Chile, the plantation of pines and other exotic conifers represents a relatively incipient but rapidly growing activity in the Argentinean Patagonia; thus, there is urgent need of basic information on its impact of on biodiversity. Our findings provide evidence on the effects that the habitat replacement associated with this activity has on three widely recognized and used bioindicator groups, also offering insights on how pine plantations could be designed and managed to reduce their impacts on biodiversity. Canopy cover seems to be a key factor related to species richness and abundance in plantations. Therefore, we suggest that more open canopies or higher heterogeneity in canopy cover, could increase biodiversity in conifer plantations in our study area. Similar results were found for ants in Patagonia (Corley et al., 2006) and Cerambycid beetles in Japan (Ohsawa, 2004). The similar, as well as different responses among assemblages emphasize the limitations and advantages of each group as a bioindicator. These limitations should be taken into account when considering the response of single groups to habitat replacement, particularly when this replacement is incipient and still occurring at small spatial scales. Our multiassemblage approach to assess biodiversity losses under habitat conversion can lead to more sound management regulations than those based on a single bioindicator group. Acknowledgments We thank M. Elgueta for the identification of the epigeal beetle species and C. Quintero for field assistance. We are grateful to T.T. Veblen, C. Bigler, V. Rusch and two anonymous reviewers for useful comments on previous versions of this manuscript. The Argentinean National Park Administration allowed us to sample in the Nahuel Huapi National Park. This study was partially funded by FONCYT (PICT 01-07320) and by the Universidad Nacional del Comahue. M.A.A. is a member of the Carrera de Investigador Cientı´fico of CONICET.

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