As the name implies, ‘alien’ or exotic plants comprise plant species not recognised as indigenous or dominating species within a particular vegetation class. Alien plants are mostly introduced from a different location into ‘natural vegetation’ of rich biodiversity and functional ecosystem. Following their introduction, alien plants become invasive by establishing their basal cover and expanding rapidly beyond current habitat boundaries (Lodge 1993). Alien species introduction are mostly products of direct continuous human re-distribution of species to support agriculture, aquaculture and recreation; indirectly through ballast water discharge, attachment to ships’ hulls and creation of new links between oceans. Approximately, 50 000 foreign species (non-native species) were introduced in the United States as a result of human population growth, rapid movement of people and alteration of the environment (Pimentel, Zuniga & Morrison 2005). There
is an increase in international trade of goods and materials among nations, and this trade patterns create opportunities for unintentional introductions (USBC 2001). Due to development and commercial activities, deliberate introductions of alien plants exist in most countries. In other words, introduction of alien species could either be accidental or deliberate; many efforts to decipher the properties that code them into becoming invasive have been attempted (Gurvich, Tecco & Díaz 2005).
In the 17th century, alien species were accidentally introduced in South Africa by the European ships that stopped at Cape of Good Hope for overhauling; but were also cultivated deliberately for commercial afforestation due to predominant slow growing indigenous trees (Beater 2006). In order to meet the increasing demand for timber, tannins, oils, firewood, wind breakers, ornamentals, charcoal etc., fast growing tree species were introduced (Nyoka 2003). One hundred and sixty-one out of about 8 750 trees introduced into South Africa (Van Wilgen et al. 2001) are regarded as invasive with 110 of the invasive species being classified as woody trees (Nyoka 2003). The introduction and purpose of early invaders is detailed by Van Wilgen et al. (2001), while emerging invaders are highlighted by Mgidi et al. (2007) and Nel et al. (2004).
Recently, growing concerns on safeguarding the socio-ecological benefits of ecosystems have overridden the initial celebration of plant transfers (Cronk & Fuller 1995); as condensed from the large body of literature pummelling on the disruptive effects of IAPs (Ehrenfeld, Kourtev & Huang 2001; Le Maître Versfeld & Chapman 2000; Dye & Poulter 1995). Plant transfers seem to be categorised as good and bad scenarios (Kull & Rangan 2008) or conflict of interest between the two, but do-nothing scenario is not sustainable because the benefit – cost ratio is not always justified (see De Wit, Crookes & Van Wilgen 2001).
Several estimates have been made (Nyoka 2003) using different methods, including concentrating on a particular specie or area, to analyse the spatial extent of alien plant invasions in South Africa (Van Wilgen, Forsyth & Le Maître 2008; Van Wilgen et al. 2001; Le Maître, Versfeld & Chapman 2000). Estimates on the range and abundance of alien plant invasions in South Africa have been made using several methods and analyses (Kotze et al. 2010; Van Wilgen, Forsyth & Le Maître 2008; Nyoka 2003). With mapped areas of 18 million hectares of IAPs in 2010 (Kotze et al. 2010), this figure has significantly increased from the rough estimates made by Le Maître, Versfeld & Chapman (2000) of only 10 million hectares in 1996 / 1997. The Southern African Plant Invaders (SAPIA) database, however, harbours IAPs data distribution records in South Africa (Lesotho) and Swaziland with a record span of 31-year period (Henderson 2010). This information database
is best used for broad-scale study (national or regional scale) for evaluation of invasion potential, degree and impacts as well as the distribution of contemporary emerging alien plants (Van Wilgen, Nel & Rouget 2007; Van Wilgen et al. 2004; Richardson & Van Wilgen 2004).
The ‘invasion paradox’ (Fridley et al. 2007) describes the successful invasion and co-existence of alien species in non-native habitat through niche differentiation (Catford et al. 2012; Lambdon Lloret & Hulme 2008) and habitat filtering (Keddy 1992). The former remarked on the unique traits of alien species over native species that promote alien invasion and dominance at different niches. Such unique traits include possession of leaves and branches (e.g. Acacia mearnsii, Lantana camara) that contains chemical inhibition of growth and seed germination of other plants (allelopathic properties) and fire-tolerance features (e.g. Cytisus spp., Brassica spp) (Richardson & Van Wilgen 2004); while the habitat filtering points out that alien species possess specific traits to pre-adapt to new habitat, probably sharing similar traits with native species. In plain terms, successful alien species invasion is dependent on alien invasiveness and community invasibility, where the capacity of alien species to invade varies significantly with ecosystem’s susceptibility. Such habitat susceptibility, for instance, accelerated the invasion of Bromus tectorum in the Great Basin Sagebrush ecosystem (Chambers et al. 2007).
Gurvich, Tecco & Díaz (2005) meta-analysed certain vegetative attributes, termed triggering attributes, from the literature to demonstrate uniqueness in trait distribution as attributes of dominance in an ecosystem. Chapin et al. (1996) proposed that the invasion of any specific species was relative to these peculiar attributes that bridged the trait distribution in a community. Such adaptive invasion by alien species through evolution was illustrated by Prentis et al. (2008). Generally, invasive intensity highly depends on the invasion history and reproduction rate of invading species (Kolar & Lodge 2001).
Conversely, Vitousek (1986) argued that the impact on biodiversity and ecosystem function was dramatically higher when an entirely new life form invaded an ecosystem. Fridley et al. (2007) confirmed the probability that rich native ecosystems serve as hotspots for alien species with associated reduction in biodiversity. Moreover, Mgidi et al. (2007) established the invasion potential of emerging invaders along with Nel et al. (2004) using climate envelope modelling (CEM). While updating the protocol for the SAPIA database, new and emerging alien plants invasions were discovered in various provinces in South Africa (Henderson 2010). For instances Senecio inaequidens and S. pterophorus, two alien plants from South Africa that invaded protected areas in Spain, were shown to establish a higher invasion rate in grassland and shrubland than in
forests between seasonal disturbances in Mediterranean plant communities (Caño, Escarré & Sans 2007).
The multivariate analysis conducted by De Gruchy, Reader & Larson (2005) strongly affirmed that the alien and native species compositions relied heavily on disturbance type, species composition and nutrient level (see also Chambers et al. 2007). Fire and grazing impacts also create favourable gradients for successful displacement of native species by alien plants in blue oaks savannah, chaparral and coniferous forests (Keeley, Lubin & Fotheringham 2003). A sub- sequential decadal mapping of fire history and invasion in Banskia woodland was found to reduce species richness and changed resource use patterns with increased level of invasion (Fisher et al. 2009). Fire, with removal of perennial herbs across elevation gradients, increases the potential for invasion and resource availability (soil water and nitrate) in sagebrush ecosystem that favour the invasiveness of Bromus tectorum (Chambers et al. 2007). Sanders et al. (2007) suggested insects as mediators in invasion dynamics.
Invader-community interaction determines the rate of displacement of native plants by alien plants (Lambrinos 2002; Troumbis, Galanidis & Kokkoris 2002) as the biotic and abiotic characteristics of a habitat can act as barriers to invasion (Levine, Adler & Yelenik 2004). There are various factors and attributes that have allowed IAPs to gain dominance in non-native habitats (e.g. Gurvich, Paula & Sandra 2005; Lambrinos 2002; Davis et al. 2000). These include the eco- physiological characteristics of invasion and invasibility (e.g. Rejmánek 1996; Roy 1990) inter alia climate change, fire regime, land use, grazing and spread vectors that interactively promote the invasion of alien invasive species (Crowl et al. 2008).
General review on invasiveness tends to greatly concur with the evidence pertaining to target- community susceptibility as compared to the attributes of alien plants (Shea & Chesson 2002). However, other factors influence invasiveness and community invasibility as found in invading species and community characteristics and dynamics (Hobbs & Humphries 1995). Such other factors include taxonomic position, seed dormancy, edge effects, resource availability and mode of reproduction (Beater, Garner & Witkowski 2008).
Grasslands, as well as savannah biomes, are noted to be extensively invaded by Acacia spp, Pinus spp, Eucalyptus spp, Salix fragilis Melia azedarach and Jacaranda mimosifolia both at riparian and non-riparian zones (Nyoka 2003). However, grasslands of the Drakensberg escarpment,
moister regions of savannah biome along the lower escarpment and the KwaZulu-Natal midlands and coastal belt seem to be worst affected (Van Wilgen et al. 2001).
Globally, most temperate grasslands are heavily impacted by agricultural reinforcement with the recent aftermath of invasion intensification. Roadside grasslands dominated by grazers and serpentine soil reveal the reverse, but contradictory effects of alien invasion as non-grazed patches close to the roadside show higher density of exotic plants than grazed patches away from the roadside (Safford & Harrison 2001). Despite the anthropogenic activities and grazing potential, there are still predispositions of fescue grasslands invasion by IAPs (Tyser & Worley 1992). For instance, Phledimus stoloniferus, an aggressive alien species, was seen to displace native species and affected agricultural activities in the invaded grasslands (Huguenin-Elie n.d).
High density stands of alien conifers were shown to reduce nearly 29 000 grassland invertebrates (composition of beetle species) at low tree density (400 trees per ha) (Pawson et al. 2010). Coleoptera diversity in Drakensberg grasslands experienced greater reduction in size in grasslands invaded by Acacia dealbata (Coetzee, Rensburg & Robertson 2007). In addition, reduced grassland ant richness and colony size was prominent in Solidago spp invaded grasslands compared to non-invaded grasslands by as it increased the workers’ foraging activities (Lenda et al. 2013). Chromoelana odorata invasion further reduced endemic fauna (spider colony) in a grassland community with high grazing intensities (Mgobozi, Somers & Dippenaar-Schoeman 2008).
Grassland invasibility might be promoted by grazing intensity, roadside deposition, resource availability and community composition (Renne, Tracy & Colonna 2006). The biotic and abiotic assemblages of grassland communities seem to be affected as a result of natural and anthropogenic tendencies (Pawson et al. 2010; Safford & Harrison 2001). Nonetheless, grassland invasibility and its associated impacts appear to be well researched and documented (Pawson et al. 2010; Lenda et al. 2013).