Ecology Letters, (2014) 17: 499–507

LETTER

John L. Maron,1* Harald Auge,2,3 Dean E. Pearson,1,4 Lotte Korell,2,5 Isabell Hensen,3,5 Katharine N. Suding6 and Claudia Stein6†

doi: 10.1111/ele.12250

Staged invasions across disparate grasslands: effects of seed provenance, consumers and disturbance on productivity and species richness Abstract Exotic plant invasions are thought to alter productivity and species richness, yet these patterns are typically correlative. Few studies have experimentally invaded sites and asked how addition of novel species influences ecosystem function and community structure and examined the role of competitors and/or consumers in mediating these patterns. We invaded disturbed and undisturbed subplots in and out of rodent exclosures with seeds of native or exotic species in grasslands in Montana, California and Germany. Seed addition enhanced aboveground biomass and species richness compared with no-seeds-added controls, with exotics having disproportionate effects on productivity compared with natives. Disturbance enhanced the effects of seed addition on productivity and species richness, whereas rodents reduced productivity, but only in Germany and California. Our results demonstrate that experimental introduction of novel species can alter ecosystem function and community structure, but that local filters such as competition and herbivory influence the magnitude of these impacts. Keywords Community assembly, exotic species, grasslands, invasion, local filters, plant competition, plant productivity, small mammals, species richness. Ecology Letters (2014) 17: 499–507

Community productivity and species richness are fundamental attributes of plant community structure. As such, there has been much work aimed at understanding the determinants of these community attributes, as well as exploring how they might influence each other (Grime 1973; Adler et al. 2011). We know that standing biomass can be affected by abiotic factors such as soil fertility and precipitation, as well as by biotic factors such as variation in local plant diversity (Rosenzweig 1968; Tilman et al. 2001; Hooper et al. 2005). Local species richness can be affected by stochastic processes such as dispersal limitation (Hubbell 2001; Turnbull et al. 2000; Zobel et al. 2000; Stein et al. 2008) as well as deterministic factors including niche partitioning, competition and consumer impacts (MacArthur & Levins 1967; Maron et al. 2012; Germain et al. 2013; Kempel et al. 2013). Interestingly, invasion of exotic species into communities can also affect community productivity and local species richness. Plant productivity is often higher in heavily invaded than nearby uninvaded sites (Wilsey & Polley 2006; Liao et al. 2008; Vil
a et al. 2011). As well, invasion can also enhance spe-

cies richness, even at small spatial scales (Stohlgren et al. 1999; Stadler et al. 2000; Sax & Gaines 2003; Sax et al. 2005). This typically occurs where many ‘weak invaders’ establish at low density; invasion can clearly reduce species richness in cases where a particular invader forms dense stands that crowds out natives (Vil
a et al. 2011). The association between invasion, increased productivity and altered species richness raises several important questions. First, what is the cause–effect relationship between invasion and increased productivity or altered species richness? Almost all studies that have documented an association between invasion and increased productivity or species richness have been correlative (but see Zavaleta & Hulvey 2004; Maron & Marler 2008). Thus, it is unclear to what extent increased productivity at invaded sites is due to invasion itself, or driven by underlying site characteristics (such as nutrient status of soils) that facilitate both invasion and increased productivity. It is also generally untested whether the increase in production at invaded sites (or changes in species richness) is due to the addition of a new species to a system or whether it is caused by the addition of new exotic species to a system. Some would argue that given similar

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INTRODUCTION

Division of Biological Sciences, University of Montana, Missoula, MT, 59812,

USA 2

Department of Community Ecology, UFZ, Helmholtz Center for Environmen-

Institute of Biology, Martin Luther University Halle-Wittenberg, D-06108,

Halle, Germany 6

Environmental Science, Policy & Management, University of California at

tal Research, D-06120, Halle, Germany

Berkeley, 130 Mulford Hall, Berkeley, CA, 94720-3114, USA

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†Present address: Biology Department, Washington University in St. Louis,

German Centre for Integrative Biodiversity Research (iDiv) Halle-Jena-Leipzig,

D-4103, Leipzig, Germany

Campus Box 1137, One Brookings Drive, St. Louis, MO 63130-4899, USA

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*Correspondence: E-mail: [email protected]

Rocky Mountain Research Station, USDA Forest Service, Missoula, MT, 59801,

USA

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circumstances, invading systems with novel native species should produce similar patterns as invading them with exotics (Davis et al. 2011). Second, what is the role of disturbance and consumer pressure in mediating the impacts of colonising species on community structure or ecosystem processes? Disturbance that removes resident species can lead to greater recruitment of added species (Turnbull et al. 2000; Zobel et al. 2000; Myers & Harms 2009; Maron et al. 2012; Kempel et al. 2013). Generalist rodent consumers, as both granivores and herbivores, can suppress the recruitment of individual species (Howe et al. 2006; Bricker et al. 2010; Pearson et al. 2011, 2012; Maron et al. 2012) and also reduce resident plant cover/productivity (Batzli & Pitelka 1970; Keesing 2000; Heard & Sax 2013). While much interest has centred on the role of generalist native consumers as sources of biotic resistance (Elton 1958; Parker et al. 2006; Liu et al. 2007; Pearson et al. 2012), standardised parallel experiments across systems that quantify the magnitude of biotic resistance to staged invasions are uncommon. In most cases we do not know how strongly native consumers suppress the productivity or diversity caused by newly arriving species. Differences in consumer pressure on natives versus exotics could, at least in part, account for the widespread correlation between invasion and enhanced productivity. Finally, how might disparate grassland systems differ in their inherent invasibility? This has been a topic of longstanding interest in invasion biology (Lonsdale 1999), yet large-scale comparisons of invasibility have been challenging because simple contrasts in the number of exotic species that different systems support, or the density of exotics in those systems, are often confounded by history and differences in propagule pressure. Addressing this question requires quantifying invasibility across systems by controlling propagule pressure and experimentally examining how in situ filters affect invasion. We staged experimental invasions across grasslands in Montana, California (USA) and Germany. These grassland systems vary in underlying productivity and soil nutrients, the composition of the generalist rodent herbivore assemblages (granivores dominant in Montana, herbivores dominant in California and Germany), background levels of invasion as well as underlying disturbance regimes (German grasslands are maintained by mowing; see Appendix S1 for a detailed comparison of these grasslands). At sites across each grassland we added seeds of 19–20 native and 19–20 exotic species (species number depending on region) to subplots that were either disturbed (to remove resident competitors) or not in and out of larger rodent exclosures. By performing seed addition experiments across such divergent grasslands we could: (1) determine the strength and consistency with which addition of novel species influences productivity and species richness, (2) assess whether these effects differ predictably depending on seed provenance (i.e. whether added species are native vs. exotic) and (3) compare the strength of local in situ filters (i.e. impacts of generalist consumers and/or resident competitors) in their ability to ‘resist’ the impacts of seed addition.

© 2014 John Wiley & Sons Ltd/CNRS

Letter

METHODS

Experimental design

We established experiments in three distinct grassland systems: (1) perennial caespitose grasslands of the Blackfoot Valley, western Montana, (2) mixed exotic, annual and native perennial-dominated, coastal-influenced grasslands on the Pepperwood Preserve, northern California, and (3) perennialdominated semi-dry grasslands in central Germany (Appendix S1). Within each region, we established experiments at 10 sites (nine sites in California), with sites separated by 1–54 km, 0.1–3.5 km and 1.6–31.6 km in Montana, California and Germany respectively. At each site within each grassland type, we randomly selected a location to establish a rodent exclosure and an adjacent paired rodent exclosure control plot of the same size. Rodent exclosures were 10 9 10 m (or at some sites 10 9 15 m) in Montana, and 5.5 9 7 m in California and Germany. All were constructed from 0.625 9 0.625 cm wire mesh fencing buried 40–60 cm deep and which extended 60 cm aboveground. Fencing was topped with metal flashing to prevent rodents from climbing over the top. We maintained snap traps within exclosures to ensure they were secure. Rodent exclosures did not exclude birds, invertebrates or ungulates, but observations (and experimental seed depots placed in exclosures in Montana and California) revealed that animals other than rodents removed few seeds (J.L. Maron, D.E. Pearson & C. Stein unpublished data) and there was little evidence of ungulate browsing in experimental plots at any site. Within each rodent exclosure and rodent exclosure control plot, we established 12 permanently marked 0.5 9 0.5 m subplots. Subplots were randomly assigned to a unique factorial combination of seed addition (native, exotic or no seeds added) and disturbance, with each treatment replicated twice (two replicates 9 +/ disturbance 9 native/exotic/control seed addition = 12 subplots). In Montana, vegetation within subplots assigned to the disturbed treatment was killed in mid-growing season in 2009 using the broad spectrum, low-persistence herbicide Roundup (Monsanto Corporation, St Louis, MO, USA). Several weeks after the herbicide application, we disturbed the top 10 cm of soil and removed dead vegetation from each subplot using a hoe. In California and Germany, no herbicide was used and instead the top 15 cm of soil was mechanically turned over and all aboveground plant parts and coarse roots were removed. We added seeds of 19 or 20 native species and 19 or 20 exotic species (depending on the region, see Appendix S2) to subplots designated for seed addition (with seed provenance randomly assigned to subplot). Added species were grassland species that occurred in the regional species pool, but were mostly uncommon locally within each region’s grassland; some species were shared across regions (Appendix S2). As added species differed in seed size, we added fewer seeds of large vs. small seeded species to account for the seed size/seed number trade-off (Haig & Westoby 1988; Appendix S2). Plant species were chosen so that within each region, the distribution of seed size, life span and affiliation to functional group

Letter

were as similar as possible between the native and the exotic species pool. Seeds were collected locally within each region. In Germany and California, seeds were added to the two replicate subplots of each treatment combination at the same time, in fall 2009. In Montana, we replicated in time in addition to space; seeds were added to one subplot of each treatment combination in fall 2009 and again to a separate replicate subplot of each treatment combination in 2010. Separate control no-seeds-added subplots of each treatment combination were also established in 2009 and 2010. We estimated the invasibility of grasslands, and how disturbance and rodents influenced invasibility, by examining how addition of native vs. exotic seeds affected the aboveground biomass, species richness and evenness in seed addition subplots. In summer 2012, 3 years after seeds were added to subplots, we scored the presence/absence of all species in each seed addition (and control) subplot (in Montana we only scored the presence/absence of added species) using a quadrat with 25 squares of 10 9 10 cm (California and Germany) or 100 squares of 5 9 5 cm (Montana), and then we harvested all aboveground biomass in each seed addition and control subplot at all sites within the three regions when vegetation was at peak biomass. In Germany, biomass and species richness data were averaged across the two replicate subplots of each treatment combination. As grasslands in Germany are managed by mowing, we simulated mowing by annual biomass harvests in and out of rodent exclosures in the years prior to 2012 as well. In California, only one replicate was harvested, and in Montana, only subplots established in 2009 (i.e. the same year the experiment was established in California and Germany) were scored and harvested. In Montana, harvested plants were sorted to species, bagged, placed in a drying over 60 °C until dry and weighed. In California and Germany, plants were not sorted by species, but bagged, dried and weighed as above. In Montana, in subplots to which we added exotic seed we clipped all flowers/seed heads from plants on which these occurred to prevent new seed input. At the end of the experiment we killed all exotics using Roundup. This was not done in Germany and California. To examine how seed addition influenced the evenness of local assemblages, in California and Germany, we visually estimated the per cent cover of all species within each seed addition and control subplot at peak biomass in summer 2012. Percentage cover of species was then used to calculate Pielou’s species evenness of each subplot using the index: J = ∑ (Pi 9 ln Pi)/ln S with Pi being the relative contribution of the ith species to total cover, and S the total number of species on subplot. Analysis

To examine how seed addition influenced aboveground plant biomass, at each site within each grassland type, we calculated a log response ratio, defined as ln(seeds added/control) where control was the total aboveground biomass of resident vegetation in the no-seeds-added subplots (of the same disturbance and rodent exclusion treatment as the seeds-added subplot used in the numerator) and seeds added was the total aboveground biomass in a seeds-added subplot of the same site/

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treatment combination as the denominator. By performing analyses on the log response ratio (rather than simply the difference in productivity between seed addition and control subplots) we could standardise across our three grassland types that varied in background productivity. We used a four-way ANOVA to determine how region (Montana, California and Germany), seed provenance (native vs. exotic seeds), rodent exclusion and disturbance (all fixed factors) and their interactions influenced the log response ratio of aboveground biomass. Site (nested within region) was treated as a random variable and rodent exclusion was treated as a whole-plot factor with disturbance and seed addition as split-plot factors. For data from California, we had to eliminate three sites from the analysis because voles breached rodent exclosures at these sites and affected vegetation inside rodent exclosures. Data were analysed in the GLIMMIX module in SAS (ver. 9.3) (SAS Institute, Cary, NC, USA) using an underlying Gaussian distribution. We used orthogonal contrasts to test a priori hypotheses and to decompose significant interactions. In particular, as the rodent community is dominated by voles in California and Germany, whereas in Montana mice are more common, we performed an a priori contrast to compare the average of California and Germany to Montana. To investigate whether adding native or exotic seeds increased productivity over the control without seed addition, we used t-tests to test whether there was a significant difference between least square means of log response ratios with zero. Any value significantly larger than zero indicates that adding seeds increases biomass. To determine how seed addition influenced total species richness, for data collected from California and Germany, we calculated the difference between total number of species on seed addition plots, and total number of species on control plots (of the same disturbance/rodent exclusion treatment combination) in 2012. This difference represents the number of newly established added species on seed addition plots. In Montana, as we only recorded the presence of added species within seed addition subplots, we calculated the difference in the total number of added species on seed addition subplots and control subplots of the same treatment combination. We then ran an identical four-way ANOVA as above, with the response variable being the difference in species richness between a seed addition subplot and control no-seeds-added subplot. A priori contrasts were constructed as discussed above. To determine how seed addition influenced the proportional change in overall species richness (i.e. resident + added species), we performed the same ANOVA as above, but using the log response ratio for species richness (defined as for biomass) as the response variable. We also ran three additional four-way ANOVAs, structured as above, but for data from Germany and California only. The response variables for these analyses were as follows: (1) the difference in the total number of non-added resident species in each seed addition and no-seeds-added control subplot (paired within site and rodent exclusion/disturbance treatment), (2) the difference in species evenness values between seed addition and no-seedsadded control subplots, again paired within site and rodent exclusion/disturbance treatments, and (3) the difference between paired seed addition and no-seeds-added control sub© 2014 John Wiley & Sons Ltd/CNRS

502 J. L. Maron et al.

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plots in the summed total percent cover of all resident vegetation (i.e. non-added species) in each subplot. These analyses allowed us to examine how seed addition affected evenness and to what extent changes in species richness and cover reflected added species vs. changes in resident community structure. Finally, to explore how particular functional groups (grasses, non–N-fixing forbs, N-fixing forbs; see Appendix S2) responded to disturbance and/or rodent exclusion, and how this differed based on seed provenance and across regions, we calculated the average cover of all added species of each functional type in Montana, California and Germany using their scores in the gridded 50 9 50 cm quadrats. In Montana, we omitted two species, Collinsia parviflora and Veronica verna because these are both early season ephemeral species that had senesced by the time we estimated cover later in the season. As some added species showed high abundances in noseed addition control plots, we corrected their cover in seed addition plots by their cover in the paired control plot. We then ran a four-way ANOVA (as above) for each of the three functional groups separately using average cover as the response variable. If a particular species did not recruit in any subplot at a particular site, that species was omitted from the data at that site.

Letter

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RESULTS

Invasibility-effects of seed addition on productivity

Across disturbance and rodent exclusion treatments, exotic seed addition led to proportionately greater productivity than did native seed addition (Fig. 1; F1,133 = 25.62, P < 0.001). As well, exotic seed addition substantially increased productivity over standing biomass in no-seeds-added plots (t = 6.42, P < 0.001 across all sites; overall effect of adding native seeds not significant: t = 1.55, P > 0.05). We calculated the difference in productivity between seeds added and no-seeds-added subplots, and compared the magnitude of this difference between exotic vs. native seed addition subplots. The difference in the effect of exotic vs. native seed addition on productivity was substantial and averaged 332.8 g/m2, 52.4 g/m2 and 45.2 g/m2 in Germany, California and Montana respectively. The log response ratio for how seed addition influenced productivity was marginally non-significantly different among regions (F2,23 = 2.7, P = 0.09). In disturbed seed addition subplots there was proportionately more biomass compared with undisturbed seed addition subplots (Fig. 1; F1,133 = 16.94, P < 0.0001). However, the magnitude of this effect also differed among our three regions (region 9 disturbance interaction, F2,133 = 4.12, P < 0.02), with California differing from Germany (contrast, t = 2.61, P < 0.01), whereas Montana did not differ from the average of the other two regions (t = 1.57, P > 0.05). In absolute terms, the difference in productivity between disturbed and undisturbed seed addition subplots (calculated as described earlier) was 217.2 g/m2, 12.4 g/m2 and 45.6 g/m2 in Germany, California and Montana respectively. Interestingly, disturbance had similar proportionate effects on productivity across both exotic and native seed addition subplots (non-significant disturbance 9 seed addition interaction; F1,133 = 2.11, © 2014 John Wiley & Sons Ltd/CNRS

–0.3 Undisturbed No rodents

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Figure 1 Effects of rodent consumers and disturbance on aboveground plant biomass in subplots to which we added native or exotic species as seed (see Methods for details). Shown are least square mean ( SEM) log response ratios for aboveground biomass, defined as ln(seeds added/ control) where control values are those from no-seeds-added control subplots of the same disturbance and rodent exclusion treatment combination as a particular seed addition subplot. Values above zero indicate that seed addition increased aboveground plant biomass compared with plots to which no seeds were added. Least square means calculated from the region 9 rodent exclusion 9 disturbance 9 seed addition interaction.

P > 0.05). Although the main effect of rodent exclusion on productivity was not significant (Fig. 1; F1,23 = 0.36, P > 0.05), the effects of these consumers varied significantly among our regions (region 9 rodent interaction, F2,23 = 5.59, P < 0.02). In Germany and California, voles were extremely abundant and vole damage to vegetation was visually striking. In contrast, this was not the case in Montana (because voles are rare and granivorous mice are more important; Maron et al. 2012). As such, the suppressive effects of rodent consumers on plant production in Germany and California were greater than effects of rodent consumers in Montana (contrast, t = 2.89, P < 0.01). The difference in plant biomass between seeds-added and no-seeds-added subplots in rodent exclosures averaged 364 g/m2, 59.12 g/m2 and 18.8 g/m2 in Germany, California and Montana respectively. In contrast, the average difference in productivity between seeds-added and no-seeds-added subplots outside of rodent exclosures was 156.4 g/m2, 4.6 g/m2 and 22.3 g/m2 in Germany, California and Montana respectively. Rodents similarly suppressed plant biomass in native vs. exotic seed addition subplots

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Experimental invasion, productivity and species richness 503

(rodent 9 seed addition interaction, F1,132 = 0.21, P > 0.05). All other two-, three- and four-way interactions were not statistically significant (Appendix S3). Invasibility-effects of seed addition on species richness and diversity

Across our three regions, adding native seeds or exotic seeds increased species richness on average by 4.7 species (t = 11.61, P < 0.001) or 5.0 species (t = 12.35, P < 0.001) respectively. There were significant differences among our regions in how seed addition influenced species richness (Fig. 2; F2,23 = 5.57, P < 0.02), with seed addition increasing species richness to a greater extent in Montana compared with Germany and California (contrast, t = 2.91, P < 0.01), and in Germany compared with California (contrast, t = 2.08, P < 0.05). Averaged across our regions, invading subplots with exotic species did not lead to greater species richness than invading with natives (Fig. 2; F1,138 = 0.67, P > 0.05), although this differed among grasslands (region 9 seed addition; F2,138 = 19.73, P < 0.0001). Exotic seed addition led to relatively greater species richness in comparison to native seed addition in Montana compared with Germany and California (contrast, t = 5.86, P < 0.001). Disturbing seed addition subplots increased species richness (Fig. 2; F1,138 = 14.23, P < 0.001), 12

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Figure 2 Effects of rodent consumers and disturbance on species richness in subplots to which we added native or exotic species as seed. Shown are least square mean ( SEM) differences in species richness between seed addition plots and plots with the same treatment that did not receive seed addition. Least square means calculated from the region 9 rodent exclusion 9 disturbance 9 seed addition interaction.

although this effect also varied among regions (disturbance 9 region; F2,138 = 7.69, P < 0.001). Disturbance had substantially greater impacts on species richness in Germany compared with California (contrast, t = 3.59, P < 0.001), whereas Montana did not differ from California and Germany combined (t = 1.57, P > 0.05). As well, there was a significant disturbance 9 region 9 seed provenance interaction (F2,138 = 6.60, P < 0.002). In Montana, disturbance greatly enhanced the species richness of exotics in comparison to natives, whereas this was not the case in Germany and California (contrast, t = 3.38, P < 0.001). Finally, excluding rodents had no significant effect on how seed addition influenced species richness (Fig. 2; F1,23 = 1.22, P > 0.05), and no other two-, three- or four-way interaction was statistically significant (Appendix S3). When we used the log response ratio for species richness as the dependent variable to examine the proportional change in species richness as a result of adding seeds, we found significant differences among regions (F2,23 = 48.84, P < 0.0001) but no other significant main effects or interactions. The addition of seeds resulted in a greater proportional increase in species richness in Montana compared with Germany and California (contrast, t = 9.81, P < 0.0001). Seed addition did not change species richness of resident non-added species in California (lsmean = 0.04, t = 1.0, P > 0.05), but slightly decreased it in Germany (lsmean = 1.15, t = 3.38, P < 0.01), with the difference between regions being significant (F1,14 = 8.16, P < 0.05). While seed provenance had no significant influence (F1,84 = 2.04, P > 0.05), the number of resident species in California and Germany was reduced in disturbed but not undisturbed subplots (undisturbed lsmean = 0.24, t = 0.73, P > 0.05, disturbed lsmean = 0.95, t = 2.93, P < 0.01; F1,84 = 12.67, P < 0.001). There were no other significant main effects or interactions in effects on resident species richness (Appendix S4). Seed addition of native species as well as of exotic species slightly increased evenness (exotic seed addition: lsmean = 0.051, t = 2.78, P < 0.01; native seed addition: lsmean = 0.056, t = 3.06, P < 0.01), in both Germany and California; however, neither the effect of seed provenance nor any other main or interaction effect on evenness was significant (Appendix S4). The total cover of resident species was not reduced by seed addition in California (lsmean = 3.4, t = 0.65, P > 0.05), but was significantly decreased in Germany (lsmean = 25.4, t = 6.27, P < 0.0001), with the difference between regions being significant (F1,14 = 11.02, P < 0.01). Disturbance decreased the total cover of resident species (F1,84 = 33.54, P < 0.0001), but these effects also differed among regions (disturbance 9 region; F1,84 = 15.62, P < 0.001), with disturbance having limited effects on total resident cover in California (lsmeans; undisturbed = 0.7, t = 0.12, P > 0.05, disturbed = 6.2, t = 1.07, P > 0.05), but decreasing resident cover in Germany (lsmeans; undisturbed = 10.9, t = 2.46, P < 0.05, disturbed = 39.9, t = 8.97, P < 0.0001). All other main effects and interactions were non-significant (Appendix S4). The average cover of added species within each of the three functional types (grasses, non–N-fixing forbs, N-fixing forbs) © 2014 John Wiley & Sons Ltd/CNRS

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turbance, F1, 134 = 10.45, P < 0.002). All other interactions were non-significant. The average cover of N-fixing forbs differed among regions (F2,20 = 8,56, P < 0.003) and was higher in California and Germany than Montana (contrast t = 2.3, P < 0.04; Fig. 3). Exotic N-fixers tended to be more abundant than native N-fixers (F1,108 = 8.16, P < 0.005), but the region 9 seed provenance interaction was significant (F1,108 = 9.83, P < 0.001) with the average cover of exotic N-fixers greater in California and Germany than in Montana (contrast t = 3.44, t < 0.0008). Finally, the effect of rodent exclusion differed across regions (region 9 rodent exclusion, F2,20 = 4.38, P < 0.03), with rodent exclusion tending to enhance the cover of N-fixers to a greater extent in California and Germany vs. Montana (contrast, t = 1.9, P = 0.07; Fig. 3).

had variable responses to disturbance and/or consumer removal. Average cover of added grass species did not change significantly across any treatment combination, and this was consistent across regions (effect of all main effects and interactions P > 0.11). The average cover of added non–N-fixing forb species was higher in California and Germany than in Montana (effect of region, F2,22 = 19.94, P < 0.0001; contrast, t = 3.57, P < 0002), higher for exotics than natives (F1,134 = 11.09, P < 0.002), greater in disturbed than undisturbed subplots (F1,134 = 127.6, P < 0.0001) and marginally non-significantly higher inside vs. outside of rodent exclosures (F1,22 = 3.72, P = 0.067; Fig. 3). As well, disturbance enhanced the average cover of added non–N-fixing forbs more in California and Germany than in Montana (region 9 disturbance, F12,134 = 32.8, P < 0.001; contrast, t = 2.44, P < 0.02), and more in Germany than in California (contrast, t = 7.25, P < 0.0001; Fig. 3). Added exotic non–N-fixing forbs had greater cover than added native non–N-fixing forbs in California and Germany compared with Montana (region 9 seed provenance, F2,134 = 4.75, P < 0.013; contrast, t = 2.69, P < 0.008; Fig. 3). Finally, added exotic non–N-fixing forbs had greater average cover in disturbed subplots than did added native non–N-fixing forbs (seed provenance 9 dis-

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DISCUSSION

Our replicated seed addition experiment in grasslands in Montana, California and Germany yielded several novel results. Perhaps most noteworthy was that invading with exotics led to substantially increased standing biomass, both over background productivity in no-seeds-added subplots as well as

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Figure 3 Effects of rodent consumers and disturbance on the average cover of native and exotic species of non-N-fixing forbs (left panels) and N-fixing forbs (right panels). Shown are least square means ( SEM) calculated from the region 9 rodent exclusion 9 disturbance 9 seed addition interaction.

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Letter

over subplots to which we added native species. Our results therefore experimentally demonstrate that patterns found in correlative studies where invasion is often associated with higher plant productivity (Liao et al. 2008; Vil
a et al. 2011) may arise because exotics themselves increase productivity. By overcoming propagule limitation through seed addition, we also increased local species richness. However, unlike for productivity, invading with natives and exotics generally had similar impacts on species richness. Competition from resident vegetation, as well as herbivory/granivory by generalist rodents, mediated how strongly added species affected community structure and productivity, with the strength of these interactions varying across grassland types. Disturbance allowed for greater final biomass and species richness in seed addition subplots. In Germany and California, these effects occurred irrespective of whether we invaded with native or exotic species. In Montana, however, disturbance increased exotic species richness more than native species richness. Exclusion of small mammals, primarily generalist voles in Germany and California, led to substantial gains in the productivity of seed addition subplots, regardless of whether the species added were native or exotic. In Montana, rodent exclusion had minimal effects on the productivity. Averaged across disturbance and rodent exclosure treatments, exotic seed addition increased productivity over background conditions (i.e. productivity of no-seeds-added control subplots) by an average of 40–163%, depending on grassland type. The increased production in exotic seed addition subplots was generally due to just a few species obtaining high abundance, but was not consistently related to affiliation to any of the plant functional groups (grasses, non–N-fixing forbs or N-fixing forbs). In Germany, exotic N-fixing species (Appendix S2) were disproportionately dominant, whereas in California, Daucus carota and Plantago lanceolata were the dominant added exotic species. Interestingly, local filters such as competition and/or disturbance did not universally favour exotic production over natives. For example, across our three grasslands exotic added species did not out produce native added species with disturbance. Thus, other factors likely contribute to the overall greater production of plots with added exotics vs. added natives or controls. Interestingly, we found no significant differences among regions in the impact of seed provenance on productivity or species richness (averaged across all treatments), indicating that the three disparate grasslands had consistently similar responses to native vs. exotic seed addition. Moreover, when all local filters were in place (i.e. when we analysed only those subplots that were undisturbed and open to rodent consumers), there were no significant differences among regions in the impacts of seed addition on the log response ratio of productivity (F2,23 = 0.03, P > 0.05) or species richness (F2,23 = 0.57, P > 0.05). Thus, no one grassland was inherently more open to invasion, so long as they were undisturbed and generalist consumers were present. Furthermore, in a separate analysis we found no consistent pattern across regions in how total per species seed weight influenced the ultimate abundance of those species (J.L. Maron, H. Auge, D.E. Pearson and C. Stein, unpublished data). The fact that seed addition in the presence of local filters increased local species richness, and did so simi-

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larly for natives vs. exotics, indicates dispersal limitation is common, and relatively equal in magnitude for the native and exotic species we added. That gains in species richness due to seed addition were similar in magnitude across grasslands (in the presence of filters) suggests congruence across systems in the magnitude of dispersal limitation. The minor impacts of seed addition on overall species evenness, the number of resident species and on their cumulative cover in Germany and California suggest that gains in both productivity and species richness due to seed addition did not come at great expense to resident vegetation. While we did not collect data to analyse these effects in Montana, we observed qualitatively similar results. Had the experiment run for a longer duration, we would expect that exotics would ultimately start to decrease resident abundance and perhaps even diversity, particularly in disturbed plots protected from rodents. Few studies have quantified how different grassland systems vary in the strength with which resident competitors or consumers inhibit establishment of colonising species, or compared how the intensity of these interactions vary depending on whether colonisers are native or exotic species. Metaanalyses that amalgamate results from different studies in divergent systems have found conflicting results. Some reviews have shown that generalist native consumers can provide biotic resistance to invaders (Parker et al. 2006; Kimbro et al. 2013). Others indicate that native plants incur greater enemy damage than exotics (Liu & Stiling 2006; Hawkes 2007), although effects of consumers on plant performance are generally similar between natives and exotics (Chun et al. 2010). Finally, other meta-analyses have indicated that the strength of biotic resistance imposed by resident competitors and consumers is roughly similar (Levine et al. 2004). Although useful, these results are often difficult to interpret because ‘invasion resistance’ is often defined differently among studies. Some studies examine effects of native consumers/competitors on the success of established invaders (i.e. Kimbro et al. 2013), whereas others examine how native species influence invader establishment or impact (i.e. Levine et al. 2004). Moreover, as meta-analyses summarise results from single experiments done in different systems, where methodology and other factors are not standardised, the generality of such results are not always clear. Using standardised parallel experiments in grasslands in Montana, California and Germany, we found that adding seeds of a diverse suite of species to disturbed subplots led to universally greater productivity and species richness compared with seed addition into undisturbed subplots that supported resident competitors. These results confirm that seed addition combined with removal of resident species often leads to greater recruitment of added species compared with seed addition alone (Gross & Werner 1982; Turnbull et al. 2000; Zobel et al. 2000; Myers & Harms 2009). Although competition by resident vegetation reduced the impacts of added species (on productivity and species richness) in all three grassland types, the magnitude of these effects also differed among systems. In western Montana, large resident bunchgrasses provided strong competitive resistance to invasion (Maron et al. 2012). Disturbing subplots in Montana allowed more added species (particularly exotics; Fig. 2) to establish than in Germany and California, partly © 2014 John Wiley & Sons Ltd/CNRS

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because resident vegetation in Montana recovers very slowly after disturbance. This slow recovery of vegetation also led to low biomass in disturbed seeds added subplots through time (although less so for subplots invaded by fast-growing exotics vs. slower growing natives; Fig. 1), which accounts for the more muted effects of disturbance on productivity in Montana vs. the other two grasslands. In Germany and California, disturbance not only facilitated the establishment of invading species, but some of these attained high biomass very quickly. This lead to enhanced production in disturbed vs. undisturbed subplots and also in comparison to disturbed subplots in Montana. Interestingly, however, resident vegetation appeared to ‘resist’ invasion similarly for exotics and natives (at least in Germany and in California; in Montana, exotics were favoured more than natives by disturbance). Generalist rodent consumers also provided substantial resistance to ‘invasion’, but only in Germany and California, where abundant voles had large but similar effects on both native and exotic biomass. However, vole impacts, along with pocket gophers in California, were not completely indiscriminant. In California, species with belowground structures such as bulbs (e.g. native Sisyrinchum bellum) or large taproots (exotic Daucus carota) were particularly reduced, whereas in Germany, voles decimated exotic N-fixers (e.g. exotic Medicago 9 varia, Onobrychis viciifolia and Vicia villosa). At Montana sites, voles are rare (Maron et al. 2010), and although deer mouse seed predation can be intense, it appears to have greater effects on abundance of specific species, particularly large-seeded ones (Bricker et al. 2010; Pearson et al. 2011; Maron et al. 2012), rather than on total aboveground plant biomass. Ground squirrel herbivory on added species was negligible (J.L. Maron & D.E. Pearson personal observations). These results, together with a lack of rodent 9 seed provenance interaction for any of the plant functional groups, suggest that feeding preferences of generalist native herbivores and granivores in the three regions are not driven by plant provenance, but by other plant characteristics such as defence and nutritional value. Our experiment provides significant insights regarding how in situ filters, theorised to control community composition and ecosystem function (Weiher & Keddy 1999), actually apply to both native and exotic plants across multiple systems. When all filters were in place, exotics had greater effects than natives, but the differences in effects were limited. However, reducing competition or herbivory resulted in gains in productivity and species richness, with the magnitude of these effects varying depending on grassland. The fact that ‘invasion’ by exotics had larger impacts on productivity and (at least in Montana) species richness than was produced through ‘invasion’ by natives suggests that seed provenance does matter, at least when local filters are disrupted. The weak effect of seed provenance when filters were in place is consistent with the notion that exotics are not inherently benefitting over natives due to release from their specialist natural enemies. However, when we experimentally disrupted local filters, it exaggerated the effects of exotics on productivity, indicating that release from generalist natural enemies (rodent consumers) and native competitors both enhance exotic productivity. Yet there was weak evidence for additive or synergistic effects of these two factors (sensu Blumenthal 2006). So why do the © 2014 John Wiley & Sons Ltd/CNRS

Letter

impacts of exotics increase more than natives from disruption of in situ filters? One possibility is that exotics are not a random sample of introduced species because immigration filters select for species that have inherently greater productivity (Van Kleunen et al. 2010), which can be expressed once local filters are relaxed. Another possibility is that exotics alter belowground processes in ways that enhance their own production, whereas this is less the case for natives. Examining these mechanisms in more detail, and exploring the relative influence of in situ filters in affecting native vs. introduced species could yield rich dividends for understanding invasion biology and community ecology. ACKNOWLEDGEMENTS

We thank the many people who helped with field/lab assistance in Montana, Germany and California, who are too numerous to mention individually here. We greatly appreciate the Montana Department of Fish Wildlife and Parks, the U.S. Fish and Wildlife Service, Pepperwood Preserve, the University of Montana and numerous private land owners in Montana and Germany for allowing us to work on their lands. This research was supported with grants from the USDA Cooperative State Research, Education and Extension Service (2005-35101-16040 to JLM/DEP, 2006-01350 to KS), NSF (DEB-0614406 to JLM/DEP, DEB-1001807 to KNS), the German Academic Exchange Service (DAAD 50750649) to HA, the Graduate School HIGRADE to LK and by generous travel support from the Global Invasions Network NSF (RCN DEB-0541673; PIs R. Hufbauer and M. Torchin). STATEMENT OF AUTHORSHIP

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Additional Supporting Information may be downloaded via the online version of this article at Wiley Online Library (www.ecologyletters.com).

Editor, Jonathan Chase Manuscript received 13 August 2013 First decision made 17 September 2013 Second decision made 5 December 2013 Manuscript accepted 24 December 2013

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