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Predicting Invasions of Nonindigenous Plants and Plant Pests
ment of immigrant ladybird beetles Coccinella septempunctata and Harmonia axyridis appears to be associated with declining populations of some native ladybird beetles and potentially will alter predatory communities in forest-agriculture interfaces (Colunga-Garcia and Gage 1998, Howarth 2000). Invasion by the fire ant Solenopsis invicta Buren has dramatically reduced the native ant fauna. In a detailed study in Texas, species richness of ants in infested areas decreased by 70%, and the number of native individuals dropped by 90%. Competitive displacement appears to be the primary mechanism. Similarly, overall non-ant arthropod diversity was reduced by 30%, and the numbers of individuals by 70% (Porter and Savignano 1990). The fire ants excluded some native species from the invaded areas, but it is noteworthy that the natives persisted in nearby uninvaded areas and that no extinctions were observed.
Predictions of loss of regional biodiversity accompanying plant invasions have been based on observations of diversity decreasing with the extent of an invasion. For example, the impact of nonindigenous plants on native plants has been documented for fynbos vegetation on the Cape Peninsula in South Africa. This ecosystem supports 2285 native plant species (Trinder-Smith et al. 1996), including 90 endemic taxa (comprising species, subspecies and varieties) (Richardson et al. 1996). Richardson et al. (1989) showed that invaded sites of the fynbos biome have fewer than half the plant species of matched uninvaded sites. Holmes and Cowling (1997) provided similar evidence: invaded sites had 60-86% fewer plant species. Mimosa pigra in northern Australia converted hundreds of thousands of hectares of open sedge wetland to shrubland, and native plants and animals were lost (Lonsdale 1993, Braithwaite et al. 1989). The Brazilian peppertree (Schinus terebinthifolius) was introduced to Florida in the late 19th century. It became widespread in the early 1960s and today is established on over 280,000 hectares in south Florida, often in dense stands that exclude all other vegetation (Schmitz et al. 1997).
Some functional groups are sensitive to the presence of nonindigenous plants, and others are remarkably resilient (Holmes and Cowling 1997). The chestnut blight fungus arrived in New York City in the late 19th century on nursery stock from Asia and in less than 50 years had spread over 90 million hectares of the eastern United States, destroying virtually every American chestnut tree (Castanea dentata). Because chestnut had made up one-fourth or more of the canopy of tall trees in many forests, the effects on the entire ecosystem might have initially been thought to be staggering (Roane et al. 1986). But other species (Quercus and Carya spp.) replaced chestnut in the canopy, leaving few notable changes in some ecosystem characteristics, such as primary productivity and hydrology, despite striking changes in other attributes, including the structure and dynamics of food webs and the social, cultural, and economic life of people (Youngs 2000). The simplification of ecological communities might make them more vulnerable to invasion (Levine 2000) or render them less stable or predictable in species composition (Tilman 2000). In extreme cases, invasive species may so reduce