Many theoretical and experimental studies have demonstrated that diversity plays a key role in determining community response to external pressures (Hooper et al., 2005; Ives and Carpenter, 2007). An important factor known to determine the invasibility of a community is the diversity of resident species which resist the invasion by alien species- a concept referred to as biotic resistance. This property of a community can arise from any effect of resident species on colonizing invaders, including competition, predation, herbivory and disease (Levine et al., 2004).
Among these competition from resident species has been studied in detail (Levine 2000; Seabloom et al. 2003). In combination with biotic factors, abiotic factors such as high temperature or salinity can also make an ecosystem resistant to invasion (Levine et al., 2004). One explanation for diverse communities being resistant to invasion is that in diverse assemblages resources remain utilized more fully and consequently little niche space remains available for potential colonists (Levine and D’Antonio. 1999).
In recently disturbed environments in which the resident vegetation is not likely to be sequestering all available resources, and in which establishment by new species is common, any factor(s) that increase the availability of limiting resources is likely to increase the vulnerability of a community to invasion as fluctuation in available resources is likely to be a strong determinant of invasibility (Davis et al. 2000). Invasion can be promoted by declines in local native species richness due to an increase in resource availability (Tilman, 1997; Naeem et al., 2000; Fargione and Tilman, 2005).
Research has shown that the relationship between diversity and invasibility may be both positive as well as negative as declines in local native species richness resulting in greater availability of resources for invaders have been shown to promote invasion (Dukes, 2002; Fargione and Tilman, 2005). Most of the studies evaluating diversity–invasibility relationship have kept extrinsic factors co-varying with diversity in natural systems constant and as such their applicability has been questioned (Fridley et al., 2007). However studies where extrinsic factors are allowed to co-vary with diversity have revealed that high diverse communities tend to be more heavily invaded, particularly at broader spatial scales (Stohlgren et al., 1999; Tilman, 2004).
Logically increase in resource availability can promote both native diversity and invasion (Huston 2004), potentially overriding the buffering effect of diversity on invasion that is evident when resource levels are held constant. Factors that increase the availability of limiting resources are known to increase the vulnerability of a community to invasion as Grime’s triangular model of plant strategies (Grime 1974, 1988) suggests that competition is less important in recently disturbed environments in which the resident vegetation is not likely to be sequestering all available resources, and in which establishment by new species is common.
Disturbance is a common force that can increase resource availability, thereby facilitating invasion (Hobbs and Huenneke, 1992; Burke and Grime, 1996; Hierro et al., 2006), potentially across all diversity levels (Davis et al., 2000; Wardle, 2001). A disturbance, such as fire, for example, can remove litter accumulation and allow for increased invader growth (Knapp and Seastedt, 1986; Baruch and Bilbao, 1999). More efficient resource capture by the invaders over natives (Funk and Vitousek, 2007) may also favour invader growth after a disturbance (Huston, 2004). Hence, by increasing resource availability, disturbance may shift the diversity–invasibility relationship evident in experimental settings from negative to neutral (Davis et al., 2000; Wardle, 2001).
That is, disturbance may increase resource availability to a point where, even at high diversity, resources are no longer limiting to invaders, as seen in studies where the influence of resource availability on invasibility swamps diversity-based resistance to invasion (e.g., Davis et al., 2000). Similarly, studies show that diversity confers resilience on communities faced with disturbance (Tilman and Downing, 1994; Hector, et al. 2010; van Ruijven and Berendse, 2010), but it is unclear whether this relationship might hold true when communities are also faced with invasion.
While many studies have examined various interactions between diversity, disturbance, and invader success (Hobbs and Huenneke, 1992), a key question that has received little attention is whether local-scale diversity continues to impede invader success even after disturbance or, by contrast, whether disturbance creates an invasion ‘‘window’’ (Johnstone, 1986) that promotes invader success equally across diversity levels. For addressing this relationship, two observational studies have been carried out.
One study in a forest ecosystem revealed that canopy disturbance was a more important predictor of invasibility than was diversity (Eschtruth and Battles, 2009), but it did not examine the interactive effects of diversity, invasion, and disturbance on the native community. Another study in a grassland ecosystem found that highly productive areas with high diversity were more heavily invaded after wildfire than were low productivity areas with low diversity (Harrison et al., 2003). However given that extrinsic factors, such as propagule pressure and resource gradients, were not controlled in these studies it is unclear whether diversity drove the observed invasibility patterns. Furthermore, as with the majority of diversity–invasibility studies, these examples examined invader success, yet did not explicitly examine impacts of invaders and disturbance on the native community (Levine et al., 2003).
Studies have shown that functional diversity, or the value and range of species traits (Power et al., 1996; Hooper and Vitousek, 1997; Mack and D’Antonio, 1998) is considered to be more important than number of species per se in determining ecosystem functioning and as such determine ecosystems response to invasion. In general, functionally diverse plant assemblages resist invasion better than less diverse assemblages (Elton, 1958; Burke and Grime, 1996; Levine and D’Antonio, 1999) and the possible mechanism of resistance include resource preemption (Davis and Pelsor, 2001; Dukes, 2001). Greater functional group diversity means greater niche complementarity and niche differentiation which is known to increase productivity and decrease community invasibility (Tilman et al., 1997; Brown, 1998; Carpinelli, 2001).
Given the importance of functional trait diversity it is predicted that modifications to within- and between community trait compositions will cause severe impact upon community and ecosystem function, and resistance to environmental change (Olden et al., 2004). Presently global change, on one hand, is threatening survival of specialists but at the same time is facilitating spread of generalist species (Clave et al., 2011; Mckinney and Lockwood, 1999). This decrease in functional diversity results in functional homogenization (Fisher and Owens 2004) and it adversely affects community structure, ecosystem functioning (Tilman et al., 1997), stability (Sankaran and McNaughton, 1999) and resistance to environmental change by simply narrowing the available range of species specific responses (Stachowicz et al., 2002).
Communities with much greater breadth in functional space exhibit higher resistance or resilience when compared with homogenized communities.Despite its importance, and the fact that species diversity is often an inadequate surrogate, functional diversity has been studied in relatively few cases. Approaches based on species richness on the one hand, and on functional traits and types on the other, have been extremely productive in recent years, but attempts to connect their findings have been rare. Cross fertilization between these two approaches is a promising way of gaining mechanistic insight into the links between plant diversity and ecosystem processes and contributing to practical management for the conservation of diversity and ecosystem services.
Anthropogenic activities, plant invasion and biotic homogenizationAnthropogenic activities have weakened biogeographical barriers to dispersal resulting in biotic homogenization on account of the global spread and establishment of alien species (La Sorte et al., 2007). Recent studies have shown that functional, taxonomic and genetic homogenization is accompanied by habitat loss for specialised species (Foley, 2005; Meyer et al., 2013). Habitat modification and species invasion are considered to be two among many inter-related and interacting global change drivers (Didham et al., 2007). Anthropogenic activities like habitat degradation, habitat fragmentation, and intentional/unintentional introduction of floral elements have caused large scale reshuffling of floras both at local as well as global scale (Buhk et al., 2017).
Simulation studies have shown habitat type is an important determinant of invasion (Barlow and Kean, 2004) and it is suggested that even a comparatively small increase in habitat change over time can lead to an abrupt increase in invader abundance e.g. habitat destruction lead to a sudden shift from native Mytilaster-dominated to invasive Brachidontes-dominated mussel communities in the Mediterranean Sea (Rilov et al., 2004).
As such human mediated habitat transformation has lead to loss of biotic distinctiveness. In addition, habitat modification can further aggravate the sublethal effects of invasive species (Suarez and Case, 2002; Ghazoul, 2004) e.g. invasion of flowering Siam weed Chromolaena odorata into logged forests in Thailand attracted butterfly pollinators away from native Dipterocarpus obtusifolius flowers, leading to a seven- to eight-fold reduction in pollinator visitation in modified habitats (Ghazoul, 2004).Gamma diversity, in any area, is strongly determined by change in species composition between different habitats (beta diversity) within the landscape (Jurasinski et al., 2009; Tscharntke et al., 2012).
This dissimilarity between habitats is in turn maintained by heterogeneity in habitat characteristics and it decreases as the habitats become homogenized either due to habitat modification such as intense agricultural use (Benton et al., 2003; Gámez-Virués et al., 2015; Buhk et al. 2017) or any other environmental alteration which facilitates invasion of widespread species and extinction of unique endemic species (Olden, 2006). This dissimilarity is also determined by geographical distance between habitats or patches (Nekola and White, 1999; Soininen et al., 2007) as environmental conditions tend to be similar in closer vicinity than when located further away, thus affecting species sorting according to their particular niches (Buhk et al. 2017).
Although habitat differences exist due to natural environmental gradients representing environmental heterogeneity but human mediated land-use changes such as development of roads, intense agriculture etc. weaken these natural environmental gradients creating relatively large patches of homogenously treated land (fertilized, ploughed, treated with pesticides, disturbed to same extent etc) which favours generalist species with a wide physiological amplitude. These are mostly ruderal species that are able to live and reproduce successfully under various environmental conditions (Ekroos et al., 2010) making environmental gradients less important for species sorting processes (Clavel et al., 2011).
Secondly, the reduction of dispersal limitations may happen because dispersal barriers like hedges disappear from the intensively used landscape (Benton et al., 2003; Ekroos et al., 2010).In modern world, the major cause of habitat loss is urbanization (Mckinney, 2002; Czech et al., 2000) and it is known to drive greatest local extinction rates which frequently eliminates the large majority of native species ( Kowarik 1995, Marzluff 2001) resulting in greatest form of biotic homogenization. In particular, urbanization often creates habitats best suited for ruderal species (McKinney, 2006; La Sorte et al., 2007; Ricotta et al. 2014) and as such cause biotic homogenization.
One reason for this homogenization is the exceptionally uniform nature of urban areas. For example, cities are habitats constructed almost exclusively to meet the relatively narrow demands of just one species i.e. Homo sapiens, which makes cities physically very similar throughout the world with roads, skyscrapers, and residential housing in the suburbs being almost indistinguishable. Because biotic change driven by human influence is unequal (Chapin et al., 2000; Sala et al., 2000), it is expected that different regions will exhibit different rates of biotic homogenization, or even different trends (Rejmánek, 2000; Marchetti et al., 2001, 2006).
Urbanization alters biodiversity in many ways (Wittig, 1991; Collins et al., 2000; Pickett et al., 2001), e.g. by altering quality of air, water, and soil (Sukopp and Starfinger, 1999), temperature regime and rainfall patterns (Landsberg, 1981; Oke, 1982), habitat fragmentation and disturbance (Kowarik, 1995). Although urbanization threatens biodiversity due to habitat destruction and fragmentation (Thompson and Jones, 1999; McKinney, 2004a; Liu et al., 2003), cities are richer in plant species than surrounding areas (Klotz, 1990; Pyšek, 1993, 1998a; Kowarik, 1995; Blair, 2001; McKinney, 2002; Araújo, 2003) but homogenous in nature (McKinney, 2004b, 2002; Kühn et al., 2004).
Biotic homogenization leads to the paradox of gaining species but losing diversity (Dar and Reshi, 2014) and it poses a serious ecological challenge in the current era as it involves species introduction and extinction. Hence the negative consequences of both processes i.e., species invasion and extinction apply to the process of homogenization (Tilman, 1999, Mack et al., 2000). Biotic homogenization acts on ecosystems in many different ways. The negative consequences of species invasion like alteration of hydrologic regimes, competition with native species etc and the resulting species extinction are all associated with this process. It alters food-web structure causing cascading effects in an ecosystem and makes communities more susceptible to species invasions. It disrupts operating food chains and food webs and as such has cascading impact on the resident native species. In particular functional homogenization increases the vulnerability of ecosystems to future attacks by alien species.
Genetic homogenization, on the other hand, compromises the ability of species to adapt to changing environmental alterations (Allendorf et al., 2001). Hence biotic homogenization acts on biodiversity at all possible levels i.e. ecosystem, species and gene level.