Competition is a central component of many of the fundamental ecological theories that have guided plant ecologists (e.g. Connell & Slatyer, 1977; Grime, 1979; Tilman, 1982), and is widely acknowledged as a principal factor determining the diversity and relative dominance of species within plant communities. Competition is commonly cited as one of the processes determining the response of plant communities to environmental change drivers such as N deposition, invasive alien species and climate change.
For example, nutrient availability can be a critical factor in determining the composition of plant communities (e.g. Gough et al., 2000; Crawley et al., 2005), and changes in nutrient availability have been shown to have radical effects on the species composition of vegetation (Bobbink et al., 1998), not least because of their impact on the relative competitive abilities of plants. Several environmental change drivers act to alter the nutrient status of plants and therefore their relative competitive abilities. Low N availability, in particular, frequently limits plant growth and the competitive ability of potentially dominant species, especially in oligotrophic or mesotrophic ecosystems (Bobbink et al., 1998), or cold systems such as arctic or alpine environments where low temperatures restrict soil decomposition processes (Jonasson, 1983; Chapin & Shaver, 1985; Nadelhoffer et al., 1991; Robinson et al., 1995).
Atmospheric N deposition can lead to increased N availability in the soil (Bobbink, 1998; Lee & Caporn, 1998) and increased dominance of competitive plant species. This in turn can lead to species loss, with plants of higher growth rate and taller stature shading other plants out of a system (Bobbink, 1998; Lee & Caporn, 1998; Van den Berg et al., 2005) or reducing the availability of water during critical stages of the growing season (Zavaleta et al., 2003). Similar responses have been found in experimental warming and nutrient addition studies of arctic and alpine environments. The more competitive growth forms, for example graminoids, are particularly responsive to enhanced temperatures or nutrient addition treatments (Zhang & Welker, 1996; Dormann & Woodin, 2002; Brooker & Van der Wal, 2003; Bret-Harte et al., 2004), probably as a consequence of both enhanced soil nutrient availability and enhanced nutrient uptake capacity in tissues (Semikhatova et al., 1992; BassiriRad, 2000). For example, after five seasons of fertilization at a subarctic dwarf shrub heath site there was an 18-fold increase in the abundance of Calamagrostis lapponica (Parsons et al., 1995; Press et al., 1998). The enhanced growth of particular vascular plants, and associated increased levels of competition, have been proposed as a cause for changes in the composition of communities (Harte & Shaw, 1995), such as the loss of cryptogam species (Cornelissen et al., 2001).
In the case of atmospheric N deposition, the impact on diversity can depend upon the initial availability of N. In systems where N availability is particularly low, increased availability can lead to increased diversity, whereas in systems where N availability is already at intermediate levels increased N can lead to the competitive dominance of particular species (e.g. Brachypodium pinnatum in chalk grasslands; Bobbink, 1991) and thus a loss of diversity from the community (as proposed in some diversity-productivity hypotheses; Grime, 1973; Huston, 1979). Similarly, in arctic and alpine environments increased temperature, whether acting directly or through changes in soil nutrient availability, can lead to increased diversity (at least within the vascular plant community) if new species from warmer environments are capable of migrating into a region (ACIA, 2004). One of the clearest impacts of increased global temperatures has been altered phenology in many species. (Root et al., 2003). This has been particularly evident in those biological events that occur at the start of the growing season, for example bud burst and flowering in plants, which now occur earlier in the year for many species (Sparks & Menzel, 2002). However, the response of species is not uniform, i.e. changes in phenology are species-specific (Peñuelas & Filella, 2001; Walther, 2003). For example, Fitter & Fitter (2002) found a particularly large acceleration in the phenology of early flowering species, and that phenological responses were dependent on the position of individuals within their species’ range (with individuals at range margins responding less markedly). Species-specific phenological changes have important implications for the interactions between plants. Early season variation in the capacity to acquire resources (which will be influenced by phenology) may, as a consequence of the subsequent impacts of competition, strongly influence the relative abundance of species within communities. Dunnett & Grime (1999) proposed that this effect of competition would be stronger in productive environments, for example grassland systems (for reasons explained in section IV of this review). They demonstrated that, when grown in monocultures, five common UK roadside species all responded positively to increased spring warming. However, when grown in mixtures the benefits of warming were only observable in a subset of species, indicating that interspecific competition had regulated the response of plant community composition to climatic fluctuations.
Although commonly thought of in the context of being a major driver of global warming, elevated CO2 concentrations can also have direct impacts on relative competitive abilities. When there are adequate supplies of other resources, increased CO2 concentrations can enhance photosynthesis in C3 plants and increase plant water use efficiency, thereby stimulating plant growth (Bazzaz, 1990). However, as with the response of phenology to climate change, these effects vary among species (and even among different populations of the same species; Bazzaz, 1990), and this differential stimulation of growth may alter the balance of plant–plant interactions. For example, Hättenschwiler & Körner (2003) found species-specific responses to elevated CO2 in the tree species of Swiss temperate forests: nonnative Prunus laurocerasus was found to respond positively to elevated CO2 concentrations, whereas some native species did not. This may alter the competitive relationships among tree species and promote the invasion of Prunus. That changes in interactions can scale up to substantial changes in community and ecosystem properties is demonstrated by the enhanced growth and biomass of lianas in tropical forests under elevated CO2, which could promote faster tree turnover and a general shift in the demographic structure of tropical forests from late to early successional stages (Körner, 2004; Phillips et al., 2004).
Hättenschwiler & Körner's (2003) study of Swiss temperate forests demonstrates the potential interacting effects of environmental change drivers. In this instance, elevated CO2 concentrations promote the success of an alien species, and alien species are themselves considered a driver of environmental change. Changes in the relative competitive abilities of native and alien species may also result from enhanced N availability, and the impact of N deposition on interactions may promote the influx of alien species into communities (Brooks, 2003). Even in the absence of altered nutrient availability or elevated CO2 concentrations, competition is an important part of the processes associated with the influx of alien species into communities. In many instances, the influx of alien species has negative effects on the existing plant community, leading to reduced diversity and the loss of native species (Levine et al., 2003). Competitive plant–plant interactions commonly play a central role in these impacts. In the native dry forest ecosystems of Hawaii, the dense roots and shoots of invading grass species negatively affect nutrient and water acquisition and germination of native woodland species (D’Antonio & Vitousek, 1992; Cabin et al., 2002), whilst in Californian coastal chaparral communities the invasive Carpobrotus edulis reduces soil water availability to native shrubs, negatively affecting their growth and reproduction (D’Antonio & Mahall, 1991). In both these cases the type of interaction is one that the native species will have experienced before, i.e. diffuse competition for resources such as water or nutrients. However, in some instances the interaction mechanism is entirely novel. For example, Centaurea diffusa (the Eurasian forb, diffuse knapweed) invades and dominates native bunchgrass communities of North America through the allelopathic effects of root exudates. These have a far greater effect on grass species from North America than on those from the native European range of C. diffusa (Callaway & Aschehoug, 2000; Callaway & Ridenour, 2004; Vivanco et al., 2004). These allelopathic effects are not competitive interactions in the commonly used sense of the term, yet they enable one species to actively dominate a community and monopolize its resources – to all intents and purposes C. diffusa is a competitive ‘winner’.
In general, experimental studies demonstrate strong competitive effects of invasive alien species on native species (Levine et al., 2003). However, it is important to realize that very few plant species introduced to communities away from their native ranges actually establish and go on to become aggressively invasive. There is now a considerable body of research examining the processes that restrict or promote invasibility in native communities, and although the nature of the relationship between species diversity and invasibility of ecosystems is still unclear (see e.g. Davis et al., 2005a; Herben, 2005), it is widely accepted that competitive plant–plant interactions can be an important component of the mechanisms that exclude potentially invasive species from communities (Shea & Chesson, 2002; Fargione et al., 2003; Davis et al., 2005a).
It is clear that competition plays an important role in mediating the impacts of a wide range of environmental change drivers (e.g. climate, N deposition and CO2 concentrations) on natural systems, and the ready evocation of competition as a mediating mechanism may reflect how well the concept of competition is imbedded in the minds of plant ecologists. More recently, however, the role of facilitative interactions in regulating the composition of communities has received increasing attention (Bertness & Callaway, 1994; Callaway, 1995; Brooker & Callaghan, 1998; Dormann & Brooker, 2002; Bruno et al., 2003). Although in some cases they can be very specific (e.g. through increased ectomycorrhizal infection of tree seedlings by conspecifics; Dickie et al., 2005), facilitative interactions are commonly diffuse. Other than with respect to a small set of commonly recognized phenomena, for example the nurse-plant effect where a mature plant provides protection and shelter for establishing seedlings in desert ecosystems (e.g. Holzapfel & Mahall, 1999; Schenk & Mahall, 2002), such processes have undergone a considerable period of neglect. However, following recent renewed interest it now appears to be generally accepted that facilitative interactions (between plants at all life-history stages) can play an important role in regulating the composition of some plant communities, and that the outcome of interactions is the net effect of both positive and negative plant–plant interactions (Holzapfel & Mahall, 1999; Pugnaire & Luque, 2001; Schenk & Mahall, 2002).
Despite the recognition of their role in regulating the success of individuals and shaping communities, there is much less evidence for facilitative plant interactions mediating the impacts of environmental change drivers. However, there are a few interesting examples and more may come to light as interest in facilitative interactions increases. For example, plant–plant interactions associated with invasive species need not necessarily be negative, nor is it always the native species that receives the impact of an interaction. Facilitative interactions can promote the survival of plant species in environmental conditions that would otherwise be too stressful (Choler et al., 2001; Cavieres et al., 2002), thus effectively expanding their realized niche (Bruno et al., 2003), and in some cases the facilitated species can be an invasive species. Cavieres et al. (2005) demonstrated how the nonnative Taraxacum officinale was facilitated at high-altitude sites in the central Chilean Andes by the cushion plant Azorella monantha. Taraxacum officinale seedling survival, net photosynthetic rates and stomatal conductance were all higher for seedlings growing within cushions than outside them, suggesting that the microclimatic modifications of the cushion facilitate the establishment and survival of T. officinale. In Brooks's (2003) investigation of the impact of N deposition on invasives in the Mojave Desert, one of the invasive species examined appeared to be facilitated by the Larrea tridentata bushes under which it grew. The positive effects of nitrogen addition were highest for Bromus madritensis beneath L. tridentata canopies, whilst they were greatest in interbush spaces for Schismus arabicus, Schismus barbatus and Erodium cicutarium. In a study of alpine treelines, Germino et al. (2002) demonstrated that improved environmental conditions as a consequence of ground-layer vegetation facilitated the establishment of tree seedlings, thus promoting the upwards shift of treeline species in response to climate change, and neighbour removal experiments (where neighbouring plants around a target individual are removed, and the response of the target is compared with that of untreated individuals to assess the net effect of neighbours on the target) in conjunction with experimental warming have indicated that both competitive and facilitative interactions play a role in the response of species to enhanced temperatures in alpine Dryas octopetala heath in southern Norway (Klanderud, 2005).
3. Evolutionary processes
When discussing the role of competition and facilitation, we are considering how plant–plant interactions regulate the diversity of existing species. However, a key process determining community composition is species evolution which, along with species extinction, sets the overall size of the global species pool from which a community is constructed. An important recent realization with respect to understanding the response of species and communities to environmental change is that species are not static entities, and can evolve rapidly in response to environmental change drivers. For example, Thomas et al. (2001) have demonstrated how insect species are evolving at range margins in such a way as to promote range expansion, and Davis et al. (2005b) have argued that the concept of rapid evolutionary responses (on the time-scale of decades to centuries depending upon the plant species involved) should be included in the analysis of paleoecological data on species range shifting in response to past climate change.
The concept of plant growth strategies, in particular the existence of a competitive strategy (as in the classic C-S-R theory, which proposed three main plant strategies – competitors, stress-tolerators and ruderals; Grime, 1974), implicitly accepts that plant–plant interactions can act as a selective force on plants. Recent studies of the impacts of environmental change drivers are also bringing to light more evidence of selective impacts of plant–plant interactions, and the possibility of coevolution in plant communities. Above, in the section on competition, I discussed the comparatively low impact of root exudates from C. diffusa on grass species from its native range, which suggests these native-range grasses have an evolved resistance to the chemicals of C. diffusa (Callaway & Aschehoug, 2002; Callaway & Ridenour, 2004; Lortie et al., 2004). Other studies have also discussed the potential for interactions to act as a selective force during the processes of environmental change. The Allee effect, the positive relationship between fitness and population size (or density) in small populations (Allee, 1931), is a mechanism by which plant–plant interactions might have a selective impact. At low densities, reduced seed set and recruitment can occur as a consequence of pollen limitation (Antonovics & Levin, 1980; Davis et al., 2004). Pollen sharing is a form of interaction between plants. The physiological characteristics of plants from an invasion front have been shown to differ from those in the source population in such a way as to overcome the negative consequences of these Allee effects, i.e. they have a greater degree of self-compatability, indicating that changes in the strength of interactions may have acted as a selective force, although it is difficult to determine whether this is the consequence of selection or founder effects (Davis, 2005).
Plant–plant interactions might also be the target rather than the driver of selection. It has been hypothesized that the release from enemies (herbivores and parasites) that a plant species experiences when transported to a new environment could lead to the evolution of increased competitive ability (the EICA hypothesis) as the selective force diverting resources to mechanisms protecting the plant from herbivory, for example phenolic compounds in leaves, would be reduced or removed (Blossey & Nötzold, 1995). However, there is currently conflicting evidence as to whether or not this process operates (Vilàet al., 2003; Maron et al., 2004b). For example, although Maron et al. (2004a) found reduced levels of foliar defence compounds from North American as opposed to native European populations of Hypericum perforatum, there was no corresponding increase in plant size or fecundity.
Overall, and in comparison to our understanding of the role of interactions in the structuring of communities, our understanding of the relationships between plant–plant interactions and evolutionary processes (with respect to environmental change drivers or otherwise) is limited. However, given increasing evidence in support of the potential for rapid evolutionary change during environmental change, the selective role of plant–plant interactions is likely to become an increasingly important research topic.
To briefly summarize, it is clear that competitive plant–plant interactions play an important role in mediating environmental change impacts, and that we are now also starting to see evidence of a similar role for facilitative plant–plant interactions and the evolutionary consequences of plant–plant interactions during environmental change. It would seem, therefore, that it is worth pursuing the research area of plant–plant interactions in relation to understanding and predicting the impacts of environmental change. Furthermore, although the environmental change drivers discussed vary widely in nature, the responses of plants to many of these drivers are governed by the same underlying, fundamental ecological processes (competition, facilitation and local adaptation). This would suggest a great deal of potential synergy between research fields dealing with environmental change drivers; for example, understanding the processes that regulate community invasibility and the expansion of invasive species could also help us to understand the phenomenon of range shifting under climate change. However, although some authors, for example Davis et al. (2005a; with respect to community assembly theory and invasion biology), have discussed the links between different fields, these research areas have not yet been well integrated.