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Exotic plant species have become conspicuous elements of ecosystems around the world (Mack et al. 2000). However, ecologists have struggled to pinpoint both the roles of biotic and abiotic drivers of invasion (Lonsdale 1999; Davis et al. 2000; Shea & Chesson 2002), and the contribution of exotic species to diversity patterns at various scales (Sax & Gaines 2003). Species richness has received most attention as a potential factor controlling community invasibility, although evidence in favour of Elton's (1958) diversity–invasion resistance hypothesis has been controversial (Levine & D’Antonio 1999; Wardle 2001). Studies emphasizing local processes of biotic resistance to invasion (see Levine et al. 2004) tend to neglect the role of large-scale factors, notably species pool sizes and physical stress/disturbance gradients, in driving patterns of exotic plant richness (Lonsdale 1999; Von Holle 2005).
A comprehensive understanding of exotic invasions may require a pluralistic approach to accommodate patterns observed at different scales. Recently, Shea & Chesson (2002) proposed a model that attempts to reconcile conflicting evidence on the relationship between invasion magnitude and native species diversity. Their model predicts an overall positive correlation between exotic and native species richness over broad spatial scales, at which extrinsic factors are expected to drive diversity gradients across different habitats (Levine & D’Antonio 1999). This pattern has been supported by observational studies (Lonsdale 1999; Stohlgren et al. 1999, 2002; Pysek et al. 2002; Brown & Peet 2003; Gilbert & Lechowicz 2005), reflecting the likely influence of dispersal processes, disturbance regimes and abiotic stress (or productivity) on exotic and native richness alike (Huston 1999). In addition, Shea & Chesson's (2002) model posits that a negative correlation between exotic and native richness may be expected over narrow ranges of environmental variation. At small scales, extrinsic factors should not change systematically and biotic resistance mechanisms such as competition and recruitment limitation would control the extent of invasion (Tilman 1997; Levine 2000; Naeem et al. 2000). In this light, species-rich communities are regarded as being more ‘saturated’ than species-poor ones (Moore et al. 2001; Stachowicz & Tilman 2005), thus offering reduced niche opportunities for the establishment of exotic species (Shea & Chesson 2002).
Specifically, correlations between overall richness measures may not adequately reflect potential interference from native residents on exotic invaders. If the exotic species pool were dominated by a particular functional group, biotic resistance would be better measured by the presence of native species with greater chances of interacting with exotics in that group (Fargione et al. 2003; Von Holle & Simberloff 2004). Thus, other descriptors reflecting potential niche overlap based on species’ functional identities (e.g. richness of specific functional groups) may be useful when seeking evidence that native diversity affects invasion success (Symstad 2000; Prieur-Richard et al. 2002; Ortega & Pearson 2005).
In this study we examine patterns of exotic and native species richness in the Flooding Pampa grasslands of Argentina. Increasing modification of native pampean grasslands over four centuries of human activity has been followed by massive invasions by alien species, which today account for c. 23% of all species in the regional herbaceous flora and have colonized all extant community types (Chaneton et al. 2002). Here, we evaluate the role of various drivers of community diversity and look for observational evidence consistent with the hypothesis that native diversity reduces invasion success (Levine & D’Antonio 1999; Shea & Chesson 2002). Although correlational analyses cannot establish causal mechanisms, they are indispensable for assessing multiscale patterns of invasion.
We use data from vegetation surveys conducted at different latitudes within the study region (Perelman et al. 2001) to test for exotic richness correlates with landscape species pools, major habitat gradients and native richness over broad and narrow ranges of environmental heterogeneity. First, we analyse changes in local plant richness across different habitat types covering the entire range of Flooding Pampa grasslands. At this scale, invasion levels would be driven by species dispersal from landscape pools and dominant abiotic gradients (Brown & Peet 2003). Secondly, we focus on native and exotic richness within grassland habitat types. At this scale, biotic interference from native species would contribute to limit invasion success (Shea & Chesson 2002). To enhance the latter analysis, we assess potential niche overlaps (or complementarity) between exotic and native species by looking at the functional composition of their respective landscape pools in each grassland type. Lastly, we examine the relationship between exotic richness and specific functional groups of native species.
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Our analysis shows that patterns of exotic vs. native richness may depend both on the range of habitat heterogeneity and the identity of functional groups involved. Native and exotic richness were positively correlated across sites when analyses encompassed the whole range of landscape heterogeneity in the Flooding Pampa grasslands. At this broad scale, covarying abiotic factors and species pool sizes seemed to constrain local community richness and invasion success (Huston 1999; Shea & Chesson 2002). When we searched for exotic–native richness relations within different grassland habitats, however, positive relationships only held for the more stressful habitats. Moreover, for one grassland type cross-site differences in exotic richness were negatively associated with the richness of a particular functional group of native species, instead of total richness. These findings indicate that not only the spatial scale of inquiry (Stohlgren et al. 1999; Brown & Peet 2003; Davies et al. 2005) but also the level of biological detail (Ortega & Pearson 2005) matters in looking for evidence consistent with diversity–invasion resistance hypotheses (Levine & D’Antonio 1999; Shea & Chesson 2002).
The overall association between richness of exotic and native species was found to be positive over a broad range of grassland habitats. This result parallels those reported for other systems (Lonsdale 1999; Stohlgren et al. 1999, 2002; Levine 2000; Pysek et al. 2002; Brown & Peet 2003; Gilbert & Lechowicz 2005). Such patterns are consistent with Shea & Chesson's (2002) model, which suggests that composite abiotic gradients may drive invasion patterns at broad scales of environmental heterogeneity. We found that native and exotic richness similarly changed along physical stress gradients associated with topsoil depth and salinity (Table 1, Fig. 4). It appears that site conditions favouring high native diversity also increase the chances for successful establishment of exotic invaders (Lonsdale 1999; Stohlgren et al. 1999; Shea & Chesson 2002). Conversely, stressful habitats would limit both native and exotic richness. This could have resulted in flood-prone meadows and halophyte steppes being less invaded than mesophyte and humid prairies. Yet, because native and exotic richness were still positively related after adjusting for the influence of habitat variables (Fig. 4d), other large-scale factors may have also driven the pattern of invasion among grassland types.
Indeed, exotic richness was higher in those grassland habitats having a larger landscape pool of exotic species, a trend also found for native richness (see Fig. 3). These patterns may be expected under a regional species-pool size limitation of local richness (Zobel 1997; Partel & Zobel 1999). Interestingly, grasslands typical of more stressful habitats contained, on average, a smaller fraction of their corresponding exotic pools than those located in the least stressful, prairie habitats. This is consistent with the proposal that species dispersal from landscape pools would be a more important determinant of invasion success at intermediate locations along major stress/fertility gradients (Huston 1999). Following this rationale, limited dispersal and/or ecological resistance (biotic or abiotic) might place additional restrictions on invasion of meadows and halophyte steppes by exotics pre-adapted to conditions in those habitats. The fine-grained mosaic of landscape heterogeneity in the study region (Batista et al. 1988; Perelman et al. 2001), coupled with cattle movement across habitats, would argue against a differential role for dispersal limitation in different grassland types. Instead, increased invasion resistance is likely to occur in meadows and halophyte steppes, which characterize physically stressful environments (Perelman et al. 2001). In these grassland types, abiotic constraints imposed by frequent flooding and elevated soil salinity may limit community invasibility (Chaneton et al. 1988; Greiner La Peyre et al. 2001; Dethier & Hacker 2005; Von Holle 2005).
Several studies have reported negative exotic–native richness correlations after reducing the spatial scale of analysis (e.g. Stohlgren et al. 1999; Brown & Peet 2003; Davies et al. 2005), a result consistent with models discussing the scale-dependency of diversity–invasibility relations (Moore et al. 2001; Shea & Chesson 2002; Byers & Noonburg 2003). Here, we did not find a negative association between the total numbers of native and exotic species per site within any of the four grassland types (Fig. 5). By focusing on each grassland habitat, and without changing the unit sample size (cf. Brown & Peet 2003), we reduced the effective range of environmental heterogeneity involved in analyses of invasion patterns across sites. However, contrary to predictions from Shea & Chesson (2002), exotic and native richness were either positively associated or showed no consistent relationship at the within-grassland habitat scale. Furthermore, the strong positive correlations observed in meadows and halophyte steppes (Fig. 5c,d) were robust to changes of biological detail incorporated in the analyses (Table 3). This finding fits the notion that in stressful habitats biotic resistance mechanisms (Tilman 1997; Levine et al. 2004) might become relatively less important in limiting exotic species establishment (Huston 1999; Dethier & Hacker 2005; Von Holle 2005). Indeed, facilitative interactions between native and exotic species should not be discarded as a factor potentially influencing diversity patterns in stressful sites (Bruno et al. 2005; Von Holle 2005).
We observed no significant association between native richness and invasion success across mesophyte prairies, the richest grassland communities in the region (Fig. 5a, Table 3). This result could reflect the frequent and relatively intense anthropogenic disturbances affecting these grasslands (Davis et al. 2000). Mesophyte prairies occupy elevated topographic positions with deep soils and are rarely affected by flooding or salinity (Batista & León 1992). As a result, they are subjected to periodic cultivation and are always grazed by livestock (León et al. 1984; Batista et al. 1988; Burkart et al. 1998). These perturbations are likely to relax competition from native perennial grasses, thus maintaining a high richness of exotic and native ruderal species commonly found in croplands and early successional old fields (Oesterheld & León 1987; Omacini et al. 1995). Patterns of exotic richness in this grassland habitat would be more likely to reflect differences in propagule pressure, and the site history of anthropogenic disturbance.
It has been argued that a more accurate interpretation of community invasibility may be gained by considering species’ functional attributes rather than total species numbers (Symstad 2000; Prieur-Richard et al. 2002; Von Holle & Simberloff 2004; Zavaleta & Hulvey 2004). Nonetheless, studies looking for correlative evidence of invasion resistance usually neglect the functional aspects of diversity (Ortega & Pearson 2005). Our results showed that native and exotic species largely belong to different functional groups (Fig. 6), with exotics being overwhelmingly represented by annual cool-season forbs. Cattle grazing has been found to promote short-lived and low-stature exotic forbs and grasses (Sala et al. 1986; Rusch & Oesterheld 1997; Jacobo et al. 2006). These functional types are poorly represented in the native flora, which is primarily made up of perennial species. Similar invasion patterns by functional group were reported for other temperate grasslands under domestic grazing, suggesting that exotics exploited novel niche opportunities created by introduced herbivores (Mack & Thompson 1982; Parker et al. 2006). In contrast, the predominance of cool-season species among the exotics cannot be attributed to their rarity in the native flora (see Fig. 6). The phenological niche occupied by alien species may in part reflect the history of invasions (Mack 1989), as most exotic herbs in the Flooding Pampas originated from temperate Europe. However, as warm-season exotics do occur frequently in cropland habitats, their rarity in grasslands might result from either competition with a native flora rich in warm-season grasses or lack of adaptation to flooding or salinity.
When we focused on different native functional groups, a clear pattern emerged consistent with the hypothesis that native diversity may limit invasion success in some habitats but not others. In humid prairies, total exotics richness and annual cool-season forb richness both decreased with increasing richness of native perennial warm-season grasses (Table 3, Fig. 7). Invasion resistance associated with high resident richness depends on the mechanisms controlling local coexistence, e.g. niche complementarity and recruitment limitation (Tilman 1997; Moore et al. 2001). As the peak growing seasons of native warm-season grasses and exotic cool-season species do not overlap, one might, in principle, assume they have complementary resource-use patterns (Fig. 8). However, the persistence of annual exotic species in these grasslands depends strongly on seedling recruitment during late summer–autumn (Oesterheld & Sala 1990; Deregibus et al. 1994; Jacobo et al. 2000, 2006). We thus hypothesize that regeneration of exotics may be negatively affected by a well-developed canopy of summer grasses (Fig. 8). Across humid prairie sites, the total cover and local richness of native warm-season grasses were directly related (r = 0.67, P < 0.0001). Humid prairies with higher numbers of warm-season grasses would present harsher microsite conditions for the seedlings of cool-season exotics like the widespread invasive L. multiflorum (Fig. 7). Native perennial grasses have been found to limit germination and survival of exotic annuals in other systems (Corbin & D’Antonio 2004). The proposed influence of warm-season grasses on community invasibility suggests that biotic resistance through native richness may critically depend on specific interactions between certain sets of native and exotic species (Fig. 8).
Figure 8. Schematic representation of phenological patterns for major native and exotic plant functional groups. EACS, exotic annual cool-season species; NPCSG, native perennial cool-season grasses; NPWSG, native perennial warm-season grasses. The vertical lines highlight two critical periods for the regeneration of common exotic species, and how they overlap with native species growth patterns. We suggest that the negative response of exotic richness to native warm-season grasses found in humid prairies chiefly reflects interference of summer grasses with seedling recruitment of exotic annuals during autumn. Gradients in native warm-season grass abundance would result from cross-site differences in grazing pressure.
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What ‘extrinsic’ factors could underpin the within-habitat pattern of native–exotic diversities found in humid prairies? Evidence suggests that grazing by domestic herbivores may indirectly drive native diversity–invasion resistance relations (Parker et al. 2006). A previous survey of humid prairies (León et al. 1984) showed that livestock grazing creates spatially explicit floristic gradients involving the replacement of native perennial grasses by low-growing annual species. Long-term exclosure studies have demonstrated that grazing drastically reduces the biomass of native grasses, while increasing the cover and richness of exotics through gap-colonization dynamics (Sala et al. 1986; Oesterheld & Sala 1990; Rusch & Oesterheld 1997; Chaneton et al. 2002; Jacobo et al. 2006). Conversely, cattle removal results in the recovery of native tall grasses and a rapid decline of exotic forbs, many of which go locally extinct. We thus contend that differences in grazing management history may generate natural gradients of native grass-species richness, and thus invasion magnitude.
In conclusion, our findings support previous claims that both physical and biotic factors operating at various scales influence community invasibility (Lonsdale 1999; Levine 2000; Naeem et al. 2000; Shea & Chesson 2002). In the Flooding Pampas, broad-scale patterns of exotic invasions were associated with habitat stress gradients, species pool sizes, and native species richness. The expected negative relationship between exotic and native richness could be detected only in one grassland habitat. Moreover, evidence suggests that the diversity of a particular group of native plants, not total richness, may provide invasion resistance. While this hypothesis remains open to experimental testing, identifying such key native functional groups may be crucial to inform management and conservation efforts.