Mutualistic effects of soil biota clearly facilitate some invasions, and some invasives alter soil-borne mutualists in ways that alter recipient plant communities (Richardson et al., 2000). Two of the strongest soil mutualisms involve mycorrhizal fungi and nitrogen-fixers, both of which improve the nutrient status of their host-plants. Therefore, interactions between these mutualists and invasives have the potential to alter soil chemistry, which in turn can affect native plant communities (Miki & Kondoh, 2002; Hawkes et al., 2005). Recent progress with molecular techniques linking microbial identity and diversity to function should dramatically improve our understanding of invasive-driven changes in microbial communities and affects on nutrient cycling (Schadt et al., 2005; Torsvik & Øvreås, 2004). Related to microbial effects, the effects of nonnative plant species on soil chemistry and ecosystem function have been described in detail by Ehrenfeld (2004) and Wolfe and Klironomos (2005).
One of the most ubiquitous mutualisms on earth is that between mycorrhizal fungi and plants. As mentioned earlier, the affect of individual mycorrhizal species can range from parasitic to mutualistic. Thus, the potential exists for new combinations of nonnative species and resident mycorrhizas to yield either strong parasitic or strong mutualistic interactions. Invasives might possibly encounter soil biota that facilitate establishment, but the potential for stronger facilitation by soil microbes in new habitats (nonnative) than in old habitats (native) does not fit into any of the current hypotheses for invasion. Here we refer to this concept as the enhanced mutualisms hypothesis (see Fig. 4c). Relative to the enemy release hypothesis (see III. Soil-borne antagonists), we know much less about how enhanced mutualisms affect invasions (Fig. 4b vs c). Furthermore, the evolutionary and ecological processes by which invading species might encounter novel yet stronger mutualists than those in the invasive's home community are not yet clear. It is not surprising for a nonnative plant to find a new mutualist partner that allows its existence in new regions of the world (Richardson et al., 2000); however, for a nonnative species to find a new mutualistic partner that drives the transformation of a species from low to super abundance does not have a clear theoretical underpinning. For haphazard encounters with mutualists from the nonnative range to drive far more beneficial relationships than mutualisms in the invasive's native range, mutualisms must be general, weakly specialized, or tend to evolve away from intense interaction strengths. Alternatively, it is important to consider that interactions between plants and mycorrhizas are not always simple two-way mutualisms where both partners benefit. For example, nonnative plants may exploit mycorrhizal associations because individual mycorrhizas form symbioses with multiple plants at one time forming a mycelial network. The nonnative species may be able to exploit the benefits of the symbiosis while escaping the mutual cost of maintaining the network. This sort of parasitism may be unstable but may help the invasive species establish and displace native species as the system reaches a new stable equilibrium (Schlaepfer et al., 2005). So far, we know little about mutualisms and invasives and to our knowledge there is no evidence for nonnative plants establishing new mutualisms that specifically lead to dominance and competitive exclusion of native species. However, there is a great deal of evidence that mutualistic interactions among soil biota and plants contribute to plant invasions.
Ectomycorrhizal fungal species tend to be more host-specific than arbuscular mycorrhizal fungi (Borowicz & Juliano, 1991), and the absence of ectomycorrhizas initially limited the introduction of many Pinus species to new regions of the world (Brisco, 1959; Poynton, 1979). However, this barrier has been overcome and appropriate ectomycorrhizas for Pinus species have been transported around the world and are now common throughout the southern and northern hemispheres (Richardson et al., 1994). There is also evidence that the invasive success of some plant species has been enhanced by the presence of native ericoid mycorrhizas (Wardle, 1991; Lazarides et al., 1997). Unlike ectomycorrhizas and ericoid mycorrhizas, arbuscular mycorrhizal (AM) fungi have been thought to neither limit nor facilitate invasion because of their cosmopolitan distribution and general lack of host-specificity (Richardson et al., 2000). Therefore, the presence of appropriate AM fungal mutualists may only allow invasion.
Mycorrhizal plants associate with large numbers of species of AM fungi (Borowicz & Juliano, 1991; Molina et al., 1992; Eom et al., 2000; Streitwolf-Engel et al., 2001). Therefore, we might not expect powerful new alliances among invasives and AM fungi that cause invasive dominance. Conversely, different partner pairings can result in highly variable ecological effects (Johnson et al., 1997; van der Heijden et al., 1998; Klironomos, 2003), creating the potential for unusual relationships and pairings in plant invasions. In the only direct experimental test of potential biogeographic effects of AM fungi, Klironomos (2002) found that the AM fungal fraction of a North American soil had only slightly more beneficial effects on rare native North American (four of five species) than invasive nonnative species (two of five species). Furthermore, these positive effects were only realized when the negative effects of soil-borne pathogens were excluded.
Despite the general lack of host-specificity in AM fungi associations, specificity in the growth responses of infected plants exist. The extreme variability in the growth responses of plants to different species of AM fungi can be a major determinant of local plant species diversity in natural systems (Johnson et al., 1997; Bever, 2002; van der Heijden et al., 2003). Klironomos (2003) tested the effect of multiple AM fungi isolates from native and nonnative sources on the mycorrhizal plant-growth responses for a number of grassland species. He found that plant growth associated with AM fungi that naturally co-occurred with a species (native AM fungi treatment) ranged from highly parasitic to highly mutualistic, depending on the combination of plant and fungal species (Johnson et al., 1997; Klironomos, 2003). Although the magnitude of responses was greater when using combinations of local plants and fungi, plant–mycorrhizal interactions varied from parasitic to mutualistic regardless of whether the source of AM fungi was native or nonnative (Klironomos, 2003). However, the nonnative AM fungi used were not necessarily associated with the tested invasive plants in their nonnative ranges. Incorporating AM fungi that are actually associated with species that either fail to establish in nonnative ranges or become problematic invasives may reveal important trends that correspond with invasive success. For example, AM fungi which have highly parasitic associations with nonnative species may repel invasives while highly mutualistic associations may facilitate invasion.
Arbuscular mycorrhizas are important mediators of competitive interactions between nonnative and native plants. Several studies have found that competitive effects of the invasive C. maculosa on the native grass F. idahoensis are mediated by AM fungi (Marler et al., 1999; Zabinski et al., 2002; Callaway et al., 2004b; Carey et al., 2004). When competing with C. maculosa, F. idahoensis plants were 171% larger when grown in field soil that was sterilized and provided with a microbial wash than when grown in field soil that was not sterilized (Marler et al., 1999). By contrast, C. maculosa grown with larger F. idahoensis were 66% larger in untreated field soil than field soil drenched with a fungicide that reduced AM fungi colonization (Marler et al., 1999). Other studies have reported similar interactions between this invasive and native grasses (Zabinski et al., 2002; Callaway et al., 2004b; Carey et al., 2004), suggesting that mycorrhizal networks mediate this interaction either through carbon transfer from Festuca to Centaurea via a shared mycorrhizal network (Carey et al., 2004) or increased phosphorus uptake (Zabinski et al., 2002). Similar general effects of soil fungi have also been shown for the annual invasive, C. melitensis (Callaway et al., 2001, 2003).
Invasive plants also affect AM fungal communities in ways that may create a plant–soil biota feedback facilitating invasion and altering native communities. In California grasslands, native species are more dependent on AM fungi than nonnative species (Vogelsang et al., 2005). The average growth response to a commercially available AM fungi species was c. 82% greater in seven native species than 10 nonnative species. This corresponded with a greater proportion of nonnative species than native species in this system occurring in families described as nonmycorrhizal. In three different experiments, they found that invaded grasslands vs neighboring areas without invasion were associated with 33%, c. 43%, and c. 83% reductions in per cent root colonization by AM fungi (Vogelsang et al., 2005) and similar findings have been reported for areas invaded by garlic mustard (Alliaria petiolata) (Roberts & Anderson, 2001; Stinson et al., 2006). In summary, it appears that invasion by nonmycorrhizal species can reduce the abundance of AM fungi, which negatively affects native plant species with strong dependencies on AM fungi. These altered soil microbial communities then facilitate additional invasion by nonmycorrhizal nonnative species, thus maintaining nonnative plant dominance and inhibiting the re-establishment of native species (‘the degraded mutualisms hypothesis’; Vogelsang et al., 2005). Others have reported that nonnative species are often less dependent on arbuscular mycorrhizal fungi (Reeves et al., 1979; Allen & Allen, 1980; Pendleton & Smith, 1983; but see Marler et al., 1999; Richardson et al., 2000; Callaway et al., 2004b). It is not clear whether the invasives inhibit mycorrhizas or preferentially invade areas inherently depauperate of mycorrhizas.
2. Nitrogen fixers
Invasive plants commonly increase levels of soil nitrogen, perhaps because many successful invasives take advantage of mutualisms with native nitrogen-fixing bacteria (Rhizobium spp. and Frankia spp.) (Allen & Allen, 1981; De Faria et al., 1989; Clawson et al., 1997; Ehrenfeld, 2003). Alternatively, invasive nonnative species may bring their symbionts with them, rather than enter into new associations with resident Rhizobium spp. (Weir et al., 2004; Chen et al., 2005). Regardless, symbioses between invasives and nitrogen-fixers are common even in areas without native flora that form these associations (Richardson et al., 2000; Weir et al., 2004). By contrast, the absence of Rhizobium inoculum or low inoculum densities can limit the invasive success of nonnative species (Parker, 2001). Parker (2001) reported that a threshold density of nitrogen-fixing bacteria is often necessary for nodule development on invading legumes. Indigenous legumes may provide necessary threshold densities (M. A. Parker, pers. comm.). Similar inoculum limitations may exist for species forming symbioses with Frankia spp. (Simonet et al., 1999).
The absence of appropriate nitrogen-fixing bacteria may limit some invasions, but Myrica faya invasion in Hawaii appears to have been highly successful because of a symbiosis with the nitrogen-fixing actinomycete Frankia (Vitousek et al., 1987; Burleigh & Dawson, 1994). We do not know if Myrica arrived with its own Frankia or if Frankia was already present in the system. Regardless, this mutualism has dramatically altered nitrogen cycling in Hawaiian ecosystems and contributed to highly altered native plant communities (Vitousek et al., 1987). Invasion by nonnative plants with nitrogen-fixing symbionts may also enhance secondary invasions by nitrophilous weedy species (Yelenik et al., 2004).
Non-nitrogen fixing invasives may take advantage of the soil legacies left by native plants and their nitrogen-fixing mutualists. Maron & Connors (1996) found that seedlings of the nonnative grass Bromus diandrus accumulated 48% more root biomass and 93% more shoot biomass when grown in soil collected under experimentally killed native nitrogen-fixing shrubs (i.e. lupines), compared with B. diandrus seedlings grown in soil collected at least 1 m away from lupines. Invasive nonnative plant species may also negatively affect the nodulation of resident plants and have direct negative effects on resident nitrogen-fixing microbes (and nitrifying bacteria) (Rice, 1964). Thus, interactions between plants and nitrogen-fixing symbionts may affect the establishment of nonnative species and their impact on resident native species.
Other microbes in the nitrogen cycle can be affected by invasives (Rice, 1964; Hawkes et al., 2005). For example, Hawkes et al. (2005) determined that invasive nonnative grasses doubled gross nitrification rates, in part by increasing the abundance and altering the composition of ammonia-oxidizing bacteria in the soil. These plant-driven changes in soil microbial communities are likely to disrupt ecosystem function and leave an invisible legacy of invasion.