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The assembly of species into communities and the importance of environmental filtering, species interactions and stochastic factors in that process is one of the central topics in plant community ecology. Several studies indicate that assembly of plant communities is mostly deterministic, governed by environmental filtering and species interactions (Weiher & Keddy 1999; Adler 2004; Turnbull et al. 2005). Other studies suggest that neutral processes act on assembly (Hubbell 2001; Watkins & Wilson 2003) and that species occurrence is mainly constrained by the ability of a species to disperse to and establish at a site (i.e. dispersal limitation; e.g. Zobel et al. 2000; Freestone & Inouye 2006; Myers & Harms 2009).
A common approach to assembly studies is to use plant functional traits (e.g. Weiher et al. 1999; McGill et al. 2006; Garnier et al. 2007; Cornwell & Ackerly 2009; Götzenberger et al. 2012). Functional traits are thought to reflect general adaptations to variation in the environment and trade-offs among different functions within a plant (Lavorel et al. 2007). Species with similar traits are assumed to occupy similar niches, have similar functional roles and respond similarly to the environment (Lavorel et al. 2007; Violle & Jiang 2009). It has also been suggested that differences in functional traits determine the outcome of competitive interactions and species sorting into communities (e.g. Keddy 1992; Díaz et al. 1998; Lavorel et al. 2007; Violle et al. 2009).
Semi-natural grasslands are among the most species-rich habitats in Northern Europe (e.g. Kull & Zobel 1991; Cousins & Eriksson 2002). These grasslands usually have a long history of grazing and mowing (Eriksson et al. 2002), and have traditionally not been ploughed or fertilized. The high species richness is thought to be an effect of large historical areas, management continuity, constant disturbance by grazing animals, low soil fertility (especially phosphorus) and propagule pressure from the region (Janssens et al. 1998; Cousins & Eriksson 2002; Eriksson et al. 2002, 2006). Because of land-use changes during the past century, species-rich grasslands have declined drastically in area and are today few and often small (Eriksson et al. 2002). Grasslands that have developed on grazed ex-arable fields often occur in the same agricultural landscape as semi-natural grasslands but are of a much younger age, at an earlier succession stage and are also comparatively species-poor (Cousins & Eriksson 2002). In contrast to semi-natural grasslands, the species composition of ex-arable fields is not linked to previous management in the area. Thus semi-natural grasslands and ex-arable fields are ideal habitats for studying community assembly, where the two grassland types represent similar ecosystems in the same landscape, but differing in historical legacy, successional stage and species composition and richness. Several studies have found that natural colonization on ex-arable fields takes a long time and may be contingent on soil fertility and distance to appropriate seed sources (e.g. Pywell et al. 2002; Cousins & Aggemyr 2008; Cousins & Lindborg 2008). However, few studies have investigated trait assembly in those communities (e.g. Öster et al. 2009a).
The overall objective of this study was to investigate how plant functional trait diversity and environmental factors influenced community assembly in species-poor ex-arable fields and species-rich semi-natural grasslands. We also explored if the difference in species richness between the two grassland types was connected to variations in those factors. We incorporated traits related to both the established phase of the plant's life cycle (specific leaf area, SLA, and leaf dry matter content, LDMC) and the establishment phase (seed mass). Species traits were also measured separately for each site, instead of assuming that species traits were constant over the environment.
For our analysis, we used among other a trait gradient analysis that decomposes species trait values into alpha (within-community) and beta (among-community) components (Ackerly & Cornwell 2007). Dispersion of traits within communities may show random distribution (e.g. Watkins & Wilson 2003), divergence or convergence (Grime 2006). If the measured traits have no influence on the probabilities of plants to disperse, survive and reproduce in a certain environment, then the trait assembly at a site should be a random sample from the regional species pool (Shipley 2009). If resource competition, and thus limiting similarity, is influencing the community assembly, we would expect to see evidence of trait divergence (e.g. Weiher & Keddy 1995; Weiher et al. 1998), as competition is stronger between species that are more similar in their resource use and functional traits (MacArthur & Levins 1967; Pacala & Tilman 1994; De Bello et al. 2009; Mayfield & Levine 2010). Therefore, only species with divergent functional traits are able to co-exist. However, some authors have suggested that if species are similar enough they can escape this rule of limiting similarity and co-exist (Scheffer & van Nes 2006; Yan et al. 2012). Trait convergence (under-dispersion) often indicates environmental filtering, as only species from the species pool with certain traits can tolerate the site conditions and are able to establish and grow there (e.g. Keddy 1992; Weiher & Keddy 1995). A complicating factor is that competition can also cause trait convergence, for example when only tall species may survive under conditions of intense competition for light (Mayfield & Levine 2010). In addition, if trait convergence caused by environmental filtering, and trait divergence caused by species interactions, are working simultaneously they could also cancel each other out, giving the impression of random assembly. However, by using an approach that decomposes species trait values into alpha (within-community) and beta (among-community) components will allow us to separate the influence of environmental filtering from species interactions. Thus, trait convergence, causing higher or lower species beta values than expected at random, would imply environmental filtering.
We also applied a fourth-corner analysis (Legendre et al. 1997; Dray & Legendre 2008) to the data, which tests the null hypothesis that species and their associated traits are randomly allocated into sites with respect to measured environmental conditions, and thus whether environmental filtering is influencing the species trait assembly. While the trait gradient analysis explores variation in trait values within and among communities to determine if environmental filtering is influencing the species assembly, the fourth-corner analysis directly links species trait assembly to measured environmental factors. Furthermore, if we see some relationship between measured environmental factors and species traits/richness/abundance, it indicates some kind of environmental filtering influencing the species assembly. If species are sorted into the two grassland types according to species traits or differences in environmental factors, we would expect to see differences in average values between the two grassland types.
Previous studies have shown that species presence and species abundance are not always influenced by the same factors, therefore the data were analysed for both weighted averages (considering species abundance) and species occurrence (considering only species presence; Cingolani et al. 2007).
The three functional traits chosen are only a fraction of all possible measurable traits, and studies have shown that different traits show different assembly patterns (e.g. Watkins & Wilson 2003; Valladares et al. 2008). Therefore, the ability of this study to demonstrate the effect of species functional trait diversity on community assembly is limited, given the small number of traits. However, the results were expected to provide an insight into the relationship between the measured functional traits and co-existence, as well as the mechanism that controls functional diversity and assembly within and between the target communities.
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We found no evidence of environmental filtering affecting the trait-based community assembly in the study systems, using the trait gradient analysis. For all traits, the range of alpha values exceeded the range of beta values, demonstrating that species varied more in trait values relative to co-occurring species than they did in the mean trait values of sites in which they occur. This indicates that most species occupied the same range of environments and therefore the underlying differences in the traits appeared as alpha trait variation. Ackerly & Cornwell (2007) suggested that if the high alpha trait differences are not associated with any mechanisms that promote co-existence, then the species are truly identical in performance (sensu Hubbell 2001). Further analyses of the species trait data and of soil factors, however, revealed that species occurrence and abundance was indeed related to both environmental factors and species traits, as well as grassland type.
Several studies focusing on many functional traits have found assembly rules only for some traits (Watkins & Wilson 2003; Valladares et al. 2008). Thus, environmental filtering might be working on other traits not measured in this study. It might also be that the traits measured are not linked to the environmental gradient among the sites. This, however, is unlikely as the fourth-corner analysis and simple correlations using the same traits revealed a relationship between environmental factors and trait assembly.
The grasslands studied cover a narrow environmental gradient, probably resulting in relatively small variation in functional traits among sites compared to other trait gradient studies (e.g. Ackerly & Cornwell 2007). Under these conditions, the trait gradient analysis might possibly have too little power to detect the influence of environmental filtering on species assembly. Fourth-corner analysis does however directly link the influence of measured environmental factors to species trait assembly and should therefore be more powerful in detecting those small differences. Similarly, comparing mean trait values between the two grassland types is probably a more successful method for revealing these differences than a method that studies community assembly among all the sites.
Other studies have also found a lack of evidence for trait-based organization of communities, e.g. for various leaf traits (Watkins & Wilson 2003), for plant height and potential above-ground biomass, and for seed mass (Schamp et al. 2008), and in a meta-analysis of 21 papers reporting results from a trait-based approach only 18% of cases reported significant deviation from the null model (Götzenberger et al. 2012). So the majority of similar studies based on traits and null models are coherent with our findings. It has, however, been implied that the failure of these studies to detect non-random assembly might simply be caused by the model design (Götzenberger et al. 2012), and this is supported by our study.
Results for other statistical methods than the trait gradient analysis revealed that species-poor sites had proportionally more species with a short generation time that produce many seeds (lower seed mass; Moles & Westoby 2004) and had higher growth rates (higher SLA; Cornelissen et al. 2003) compared to more species-rich sites. In addition, species with larger seedlings (higher seed mass; Moles & Westoby 2004) and thinner leaves (lower LDMC; Cornelissen et al. 2003) were more abundant on species-rich sites. Furthermore, both soil moisture and soil phosphorus also affected species assembly at a site. Species richness increased with decreased phosphorus levels, and fourth-corner analysis revealed that trait assembly into the grasslands was non-random in relation to both site phosphorus level and soil moisture. This begs the question if the link between species richness and measured traits is a direct link related to species seed production and/or competitive abilities, or a link mediated by the relationship between species traits and environmental factors. Other studies have shown that, on average, perennial species in nutrient-poor and/or dry habitats have thicker or tougher leaves than those occurring in more resource-rich habitats (Fonseca et al. 2000; Niinemets 2001; Wright et al. 2002; Cornwell & Ackerly 2009), and higher SLA values (thinner leaves) are often found for species in resource-rich environments (Cornelissen et al. 2003). This is only partly consistent with our results, which found that species with lower SLA (thus thicker leaves) were more common on sites with low phosphorus levels and more abundant on wetter sites. Species with higher LDMC (tougher leaves) were associated with wetter site and sites with higher phosphorus levels. These contradictory results indicate that perhaps the variation in SLA and LDMC is not directly linked to variation in environmental factors but rather mediated by some other factors. In addition, at resource-rich sites (sites with high soil phosphorus and moisture), species that mature quickly and produce many seeds were more abundant, while in resource-poor environments species with large seeds and thus higher establishment rates (Leishman & Westoby 1994) were more abundant.
To summarize, in the studied grasslands, as species richness increases, species dominance shifts from species that can quickly establish at a new site to species that produce stronger seedlings. Whether this shift is mostly related to increased competition, dispersal limitation or to changes in environmental factors, is hard to determine.
Differences between the two grassland types
There was a significant difference in species composition and species richness between the two grassland types. Semi-natural grasslands had much higher species richness, and the species composition was more similar among different semi-natural grassland sites than among ex-arable fields. Semi-natural grasslands had higher average seed mass, both for occurrence and abundance-weighted data, implying that large-seeded species were lacking in ex-arable fields. We postulate that the lack of large-seeded species in ex-arable fields is related to the younger age and earlier successional stage of these ex-arable fields. This is strengthened by the lack of many late-successional species in ex-arable fields, such as Anemone nemorosa, Betula pendula, Calluna vulgaris, Hieracium umbellatum, Vaccinium myrtillus, Scorzonera humilis, Succia pratensis and Viola palustris. In young communities, it is likely that easily dispersed species are favoured in the colonization process (Öster et al. 2009b). Species with low seed mass may be faster to establish there as they often produce more seeds and are therefore more likely to colonize the area, even though larger seeds usually have higher recruitment rates than smaller seeds in grasslands (e.g. Jakobsson & Eriksson 2000; Ozinga et al. 2005; Kahmen & Poschlod 2008). In addition, species with lower seed mass usually have a shorter generation time (Moles & Westoby 2004) and could therefore be faster to establish and to increase in abundance at young sites. Previous studies have also shown that colonization is more dependent on seed abundance than on resource content of the seed (Duarte et al. 2007). The more abundant a species is in the seed rain, the more likely it is to disperse to a site, thus overcoming the dispersal limitation. Other studies in ex-arable fields in Sweden (Öster et al. 2009a) and in the UK (Pywell et al. 2002) have also found that dispersal limitation is an important factor in explaining the low species richness found there. Dispersal is not only affected by seed abundance, but dispersal vectors, among other factors, can also influence species dispersal potentials (Cornelissen et al. 2003; Vittoz & Engler 2007), especially in fragmented landscapes. However, in our study area all species, unrelated to dispersal vector, should be able to disperse freely between the two grassland types, as the ex-arable fields in this study were all close to semi-natural grasslands, without any physical dispersal barrier between them.
When looking at traits related to the established phase of the plant life cycle, we found no differences in LDMC trait values between the two grassland types, but there was a higher occurrence of species with higher SLA in ex-arable fields, indicating that species with lower SLA were missing from ex-arable fields. Higher SLA has been linked to a higher growth rate (Cornelissen et al. 2003), which could imply that plants ‘mature’ more quickly, which is consistent with the conclusion that lower seed mass in ex-arable fields can be explained by shorter generation time of small-seeded species. Even though SLA seems to influence which species colonize ex-arable fields, it does not affect the abundance of species. This is coherent with findings from Cingolani et al. (2007), who concluded that the traits which determine the probability of species presence at a site are not necessarily the same as those which determine site abundance.
Species with higher SLA and heavier seeds often have better competitive abilities and higher establishment success (Moles & Westoby 2004), indicating that the differences observed between the two grassland types could, in addition to dispersal limitations, be caused by limitations on species establishment in ex-arable fields. If this is the main cause of the differences in species assembly between the two grassland types, it would imply that species with better competitive abilities and higher establishment success had lower establishment rates in species-poor ex-arable fields than other species. We found no differences in measured environmental factors between the two types of grassland that could explain these differences, which is consistent with other studies in the region (Öster et al. 2009a). There was, however, a tendency towards ex-arable fields having higher soil phosphorus levels than semi-natural grasslands, but the phosphorus levels were usually below 5 mg·100 g−1 soil, which has been suggested as a level at which phosphorus starts to severely reduce recruitment in grassland (Janssens et al. 1998). Phosphorus levels were negatively correlated with average seed mass. High soil phosphorus levels in ex-arable fields might thus influence which species can establish there. Furthermore, additional soil factors, not measured in this study, may differ between the two grassland types, thus influencing species establishment. Other studies have however concluded that seed limitation, rather than high soil fertility, is the main constraint on species colonization in ex-arable fields (Pywell et al. 2002).
Thus, even though we imply that differences between the two grassland types is mainly caused by dispersal limitations, we cannot exclude that the differences are also related to establishment limitations, caused by e.g. differences in environmental factors or species competition not captured in this study.