Niveles de evidencia y grados de recomendación

The impacts of IAPs can be described as the alteration of entire ecosystem processes (Mack et al. 2000). The impact of IAPs is apparent in the distortion of habitat structure through fragmentation, alteration, destruction or replacement. These impacts have rippling effects through reduction in biodiversity to disruption of ecosystem functions and services. For example, the removal of hemlock woolly adelgid (a small, aphid-like insect), had direct impacts on the structure,

composition and ecosystem function (Kizlinski et al. 2002), as well as indirectly affecting the composition and distribution, of bird populations in the community (Tingley et al. 2002

).

Description of IAPs impacts can be either theoretical or anecdotal. It can also be explained in various forms. For the present study, it has been narrowed to the ecological systems pertaining to this study.

As compared to indigenous trees, alien plants are known to extend their basal roots deeper into the soil, sucking out comparatively larger volumes of water and out-competing indigenous plants. In riparian zones, run-off and stream flows have been obstructed by self-established alien invasive stands due to increased evapotranspiration (ET) rates. For example, changes in stream flow voluminously increased in a fynbos catchment (mountainous) after clearing of riparian vegetation invaded by Acacia spp. in the Western Cape (Prinsloo & Scott 1999). Diurnal transpiration of IAPs causes stream flow fluctuations but with a cloudy moist climate, evapotranspiration decreases as evaporative air demand is lessened (Dye & Poulter 1995). Difficulties exist in establishing IAPs’ impacts on water use or water resources, consequently few studies have attempted such impact quantification of IAPs on water resources (Dye et al. 2001; Le Maître, Versfeld & Chapman 2000; Prinsloo & Scott 1999).

Alien plants’ novel strategies competitively absorb soil nutrients and moisture and alter the accessibility, timing, abundance of these resources to resident species (Ehrenfeld & Scott 2001). Because of their novelty, in terms of their biochemical and physiological composition, alien plant invasion of any plant community predispose their strong influence on soil nutrients and cycles (Ehrenfeld 2003; Ehrenfeld, Kourtev & Huang 2001; Ehrenfeld & Scott 2001). For example, C4 alien grass (Microstegium vimineum) litter and Bromus tectorum immobilised Nitrogen (N) with high carbon and nitrogen concentration at low decomposition rates (Thorpe & Callaway 2005; Ehrenfeld, Kourtev & Huang 2001).

Myrica faya promoted invasion of other exotic earthworms that depleted soil-stored nitrogen, altering the natural nutrient cycling through its elevated increase of soil nitrogen (Thorpe & Callaway 2005). Non-nitrogen fixing IAPs’ litter quality might influence the populations and activities of nitrogen-fixing bacteria with no symbiotic relationship, thus altering nitrogen input in the soil (Thorpe & Callaway 2005). For example, altering of nutrient cycles by alien plants (Centaurea maculosa) through root exudates (polyphenol) impacted strongly on some groups of bacteria and some processes of the nitrogen cycle of the soil (Bais et al. 2003). Bromus tectorum dramatically altered phosphorus cycling through its novel accessibility to recalcitrant phosphorus

unavailable to native plants, thereby increasing its dominance over the natives (Thorpe & Callaway 2005). Conclusively, IAPs exhibit novel schemes that can alter soil community, nutrient availability, nutrient cycling and existing plant-soil interaction, as well as water resource availability, that will proportionately affect ecosystem function and productivity.

A considerable amount of literature has captured the magnitude of IAPs impact on ecological, economic and social structures using various cumulative, systematic and empirical methods of assessment and evaluation (Pysek et al. 2012; Vila et al. 2011). The effects of IAPs on the delivery of ecosystem services from which all life forms are sustained is often excluded (Hulme et al. 2013). However, the global overview of significant ecological impacts of IAPs in literature provided highlights the complex interaction existing between the characteristics of IAPs and the susceptible environment (Pysek et al. 2012). Synthesising 199 articles to describe impacts of 135 alien plants taxa on ecosystem and surrounding on a global scale, it was found that of the 24 different impact types studied, 11 were noted to be significantly affected by alien plants, although the magnitude and direction differed within and between various impact types (Vila et al. 2011). Pysek et al. (2012), however, argued that there was no global measure of alien impact and that impact measurement solely depended on the existing context. Ehrenfeld (2003) conducted a similar meta- analysis on the effects of invasive alien species on soil nutrient cycling processes by reviewing literature on soil-related processes, soil moisture and comparison between exotic and native species in terms of co-occurrence and displacement. It was found that the spatial and temporal alien species impacts were not significantly varied among ecosystem types, but that the impacts to heterogeneity and consistency varies within an impact type. However, Simberloff (2014) found that ten percent of invading species have specific and obvious impacts on different components of the natural ecosystems. Nonetheless, the biases in various syntheses and individual impact assessments stemmed from the neglect on the biogeography of the study species and impact extent, that has no direct translation to ecosystem service reduction (Hulme et al. 2013), not until recently (see Petz, Glenday & Alkemade 2014).

The measurement of specie-impact assessment is often impractical and difficult due to high variability of biodiversity in an ecosystem (O’Connor & Kuyler 2009) and scarce knowledge and understanding of the ecosystem vulnerability to degradation and biodiversity loss (Cowling & Heijnis 2001; Holmes & Richardson 1999). Habitat destruction varies between species, often making it impossible to quantify the impacts at measurement scale, since the degradation threshold of an ecosystem which could lead to biodiversity extermination are not clearly established (Didham et al. 2007; Cowling & Heijnis 2001). Commonly, biodiversity measurement is factored

into spatial evolutionary and ecological processes that are difficult to quantify, but often easy to input into spatial analysis methods (Cowling & Heijnis 2001). No acceptable standards for impact measurement of land-change exist to fully quantify the devastating impacts of LULCC (Reyers et al. 2001).