The Impact of Invasions
Harmful invasions are the outcome of a series of low-probability events. Williamson’s “Tens Rule” (Williamson 1996) suggests that less than 1% of species that are introduced into a new environment will become damaging pests. Although organisms that arrive and establish themselves in a new range are positioned to have adverse effects on the surrounding flora and fauna, only a small fraction have been shown to do so, and that is true for both intentional and unintentional introductions. The ones that do become invasive can have, however, staggering economic and environmental costs. Economic costs—$137 billion/ year by one estimate (Pimentel et al. 2000)—of invasions by species of plants, animals, and microorganisms generally do not include the displacement or extinction of native species that are of no immediate economic concern or the effects on native ecosystems, such as changes in fire regimes, nutrient cycling, or hydrology. Moreover, for every class of invasions, many effects probably go undetected or unmeasured. Because there are few resources available to combat damaging invasive species, it would be ideal to identify the most threatening organisms and the most vulnerable communities.
There is no uniform agreement among investigators in this field on how to judge the severity of an impact, partly because scientists have not been able to carry out comprehensive and carefully designed studies of invasions. Probably the best published data on establishment and impact of introduced species come from work on biological control, which is particularly relevant for assessing impacts in forest, range, and agricultural ecosystems because most biological control efforts on both islands and mainlands are in these ecosystems. Few generalizations related to predicting the magnitude of an invader’s impact are
accepted (Vitousek 1990, Parker et al. 1999, Simberloff and Von Holle 1999, Goodell et al. 2000). One is that, in the aggregate, impacts (however defined) are likely to increase with the number, geographic range, and abundance of the introduced species. Some support for that prediction comes from well-studied natural enemies released for biological control, in which suppression of target plant species increases with the number of control species established (Hoffmann and Moran 1998). A second is that the more geographically isolated the biota of a region is, the more vulnerable it is to invasion. The latter generalization is based largely on the conventional wisdom that oceanic islands are more easily invaded than mainland areas and that native species on oceanic islands are somehow more vulnerable than those on the mainland (Elton 1958, Carlquist 1965, Wilson 1965). The explanation for that vulnerability is that the descendants of the native species on islands experienced new colonization events infrequently and might have evolved in the absence of competitors, predators, or pathogens (Carlquist 1974). However, Simberloff (1995) finds little support for the conventional wisdom that islands are more easily invaded than mainland areas or that ecological effects of invasions are greater on islands than in mainland situations. Failed immigrations are often unrecorded, and claims that an introduced species has displaced a native one are often based on correlated population changes rather than experiment or detailed field observations. Species transfers from mainland to islands appear to be far more common than transfers from islands to mainland. In Simberloff’s view, the data are inadequate to support conclusions about the invasibility or fragility of islands and mainland. A third generalization is that invaders likely to have impacts are those which create changes in disturbance regimes for which native species are ill prepared (Mack 1989): native species are especially likely to be adversely affected by the arrival of new immigrants because nothing in their evolution would likely have been comparable with the interactions suddenly faced.
Those generalizations do not identify specific potential invaders because they ignore mechanisms; the mechanisms of an invasion that threatens native species on islands are as varied as the characteristics of the invaders. For example, an invader might duplicate a functional role already played by resident species but be competitively superior in this role (Mooney and Cleland 2001), reducing the natives’ role and threatening their existence. Or an invader could perform novel functions in its new community, as when a mammalian predator invades an oceanic island that lacks native mammalian predators (Elton 1958) or a nitrogen-fixing plant invades a region of nitrogen-poor soil (Vitousek et al. 1987, Vitousek 1990).
Although the mechanisms of invasions of an isolated biota are too varied and too poorly characterized to be applied elsewhere, the identification of the vulnerability of an isolated biota is, from a management perspective, a reason at least to take special precautions to keep nonindigenous species from proliferating there. The determination of measures for all aspects of the impact of invasive species
would be useful in strengthening and focusing the ability to predict harm. This chapter reviews different scales of biological organization (individual, population, community, and ecosystem) in which impacts have been identified and discusses the development of methods to assess, quantify, and compare impacts.
DEFINING AND MEASURING IMPACT
There are few guidelines or widely accepted protocols for measuring the impact of an invader. Parker et al. (1999) suggest that the overall impact, I, of an invader on a geographic scale can be related to three factors: the total area occupied by the species, R; its abundance, A; and some measure, E, of the impact per individual or per unit of biomass. Impact can be represented as a linear combination of factors in the equation I = R × A × E. If a nonindigenous species is widespread and abundant, prediction and comparison of the impacts of invaders rest with forecasting the per capita effect, E.
Predicting invasions with that framework is fraught with obstacles (Parker et al. 1999). First, there are difficulties in measurement: per capita effects are harder to define and measure than abundance and range. Per capita effects resemble the coefficients linking interacting species in competition and predator-prey models (where the change in abundance of one species is linked to the abundance of another species via a coefficient symbolizing the strength of their interaction), but they are seldom measured in empirical studies. Second, R, A, and E are not independent. For example, it is well known from other contexts that local abundance is positively correlated with the area of an organism’s range (Gaston et al. 1997, Holt et al. 1997). Patterns of correlation among R, A, and E and the underlying mechanisms that connect them warrant further investigation. Third, the functional form of the relationship among R, A, and E can be nonlinear.
Nonlinearities can arise, for example, if the per-unit effect varies with the density of organisms. Fourth, cumulative effects of multiple invasive species are not addressed in the framework; interactions among multiple species can be antagonistic, independent, or synergistic. A common form of synergistic effect arises from the intimate interaction between a plant and its fruit or seed disperser. An invasive plant could be spread widely if a nonindigenous bird or other frugivore were dependent on the plant’s fruit or seeds. As the fruit-eater spread the plant, there would be more plants in new locales that could support more fruit-eaters, and so on. Synergistic interactions among invaders, in which one invader facilitates another, have been documented (Simberloff and Von Holle 1999). It is too early to claim any general patterns in the frequency of synergistic, independent, and antagonistic effects among invaders. As the numbers of invasive plants, animals, and pathogens in the United States continue to grow, their potential for interacting in expected and unexpected ways will increase.
Parker et al. (1999) review impacts of invaders at five levels of biological organization: effects on individuals (including rates of growth, development,
birth, death, and movement), genetic effects (including hybridization), population dynamic effects (mean and variance in abundance, population growth rates, and so on), community effects (species richness, diversity, and trophic structure), and effects on ecosystem processes (primary or secondary productivity, hydrology, nutrient cycling, soil development, and disturbance frequency). From a sample of the literature on the impacts of invaders, Parker et al. (1999) recorded the number of published studies that produced quantitative data at five biological levels on five guilds of invading species: freshwater fish, freshwater invertebrates, marine invertebrates, algae and vascular plants, and insects and other terrestrial invertebrates. Population-level effects were the most frequently studied impacts. Studies of community-level impacts were common only for plants. The least-studied impacts, except for freshwater fishes, were genetic changes and long-term evolutionary effects.
Invaders can have a variety of effects on the performance of individuals. For example, invasive plants can compete with native plants and reduce their growth (Gentle and Duggin 1997) and change their structure, such as rooting depth (D’Antonio and Mahall 1991). Invasive insects and pathogens can decrease hosts’ rates of growth, development, survival, reproduction, and movement. Such changes in individual growth and life cycles can translate into changes in population size and fate. Population models built around a simple description of the life cycle can be used to link the individual and the population (Caswell 2000).
Nonindigenous species sometimes invade areas inhabited by closely related native species. If nonindigenous and native species interbreed, genetic exchange between the two species can alter the genetic makeup of native populations. Such hybridization between an invader and a native is common and can have several consequences, including the spawning of new invaders (Rhymer and Simberloff 1966). Repeated hybridization of North American cordgrass (Spartina alterniflora) with British native cordgrass (S. maritima) eventually yielded a new, highly invasive species, S. anglica (Thompson 1991). Production of hybrid swarms and widespread introgression can lead to virtual extinction of native taxa through a swamping of their original genomic makeup through recombination with genes of the invader (Rhymer and Simberloff 1996), especially if the invader becomes much more common than the native species; this seemingly hypothetical threat is both real and serious: three species federally listed as endangered in the United States have gone extinct since enactment of the Endangered Species Act because of hybridization with nonindigenous species (McMillan and Wilcove 1994). Finally, hybridization among plants may spawn invasion by plant pathogens.
According to the “hybrid bridge hypothesis”, hybrid plants that are morphologically, genetically, and spatially intermediate between parent species can make it easier for pathogens to acquire new hosts (either by direct transfer to hybrids or indirect transfer to new hosts using hybrids as a “stepping stone”) (Floate and Whitham 1993).
The addition of novel, including invasive, organisms in new ranges creates opportunities for evolution and makes prediction of impacts more difficult. Adaptations arising in plants when they move to new environments can produce invasive descendants. The tropical alga Caulerpa taxifolia evolved tolerance of low temperatures through its cultivation in an outdoor tank at the Stuttgart Zoo (Meinesz 1999). Evolution in native animals in response to the entry of invaders can make it difficult to restore the web of interactions characteristic of the native community. For example, the native checkerspot butterfly Euphydrias editha evolved a change in host preference from a native to an introduced plant in response to a decline of its native host and an increase of a novel, invasive host, Plantago lanceolata (Singer et al. 1993). By the time the native host is restored to its original abundance, if ever, the insect population might have lost the capacity to use it.
Introduced biological controls can become less effective because of evolved changes in the virulence of control organisms or changes in the resistance of target organisms. Myxoma virus was released in Australia to control the introduced European rabbit Oryctolagus cuniculus. Within a few years, the virulence of the virus had declined, and the resistance of the rabbits had increased sharply (Dwyer et al. 1990). In other cases, nonindigenous biological control organisms appear to become more effective by adapting to novel hosts. An ichneumonid parasitic wasp, Bathyplectes curculionis, imported to the United States to control the alfalfa weevil, Hypera postica, was originally ineffective against the Egyptian alfalfa weevil, Hypera brunneipennis. Initial dissections showed that 35-40% of the wasp’s eggs were destroyed by the immune response of the larval weevil, whereas samples taken 15 years later showed only 5% egg loss (Salt and van den Bosch 1967).
Evolutionary adjustments of control and target organisms in biological control programs are not well known, but in documented cases in other contexts invading organisms have adapted to their new environments and organisms have adapted to new invaders (Mooney and Cleland 2001). Huey et al. (2000) demonstrated that introduction of a new fruit fly into the West Coast of North America resulted in the evolution, in only 20 years, of an apparently adaptive cline related to wing size throughout the fly’s vast new latitudinal range extending from southern California to British Columbia. The cline that developed in North American female flies was similar to that found in the European native populations. The developmental basis of the cline of wing size was different between the European flies and the invader in North America, although the functional result was the same and provided additional evidence of the adaptive advantage of this set of
traits. Cody and Overton (1996) described the reduction in dispersal ability for wind-dispersed seeds of invasive species onto islands in just a few generations in small isolated populations. Carroll and Dingle (1996) indicate that populations of the soapberry bug (Jadera haemotoloma) have evolved differing beak lengths in response to the introduction of new invasive hosts within only 50 years and Singer et al. (1993) have shown rapid evolution in the feeding preferences of the Euphydryas butterfly for the invading herb Plantago lanceolata in only 10 years. Thus, there are many cases of evolution both in invading species and in the species affected by invaders.
As noted earlier in this report, the ability of pathogens to adapt to different plant genotypes has been studied in detail, and the resulting knowledge has helped in forecasting the fate of new microbial genotypes in the environment (Mundt 1995). Fungal, bacterial, and viral plant pathogen populations evolve quickly to overcome resistance genes in hosts. For example, the average useful life of race-specific genes for resistance to fungal rusts of wheat has been estimated to be only 5 years (Mundt 1995).
Population and Community Effects
Invaders can cause reduction in the biological diversity of native species and the size of populations; next to land transformation, they are the most important cause of extinction (Vitousek et al. 1996). After habitat destruction (which affects 81% of imperiled plant species), introduced species contribute more to the imperilment of species (57%) in the United States than the next three causes combined— pollution (7%), overexploitation (10%), and disease (1%) (Wilcove et al. 1998, 2000) (Categories are nonexclusive and do not sum to 100%). Replacement of natives with nonindigenous species is immediate, readily measurable evidence of the impact of invasions.
Extinction could be the most dramatic impact of invasive species. Small populations of natives suffer the highest risk of extinction from various genetic and demographic causes discussed earlier in this report in connection with the same hazards that small immigrant populations experience. Invaders pose a major risk to threatened and endangered species: about 400 of the 958 species that are listed as threatened or endangered under the Endangered Species Act are considered to be at risk primarily because of competition with or predation by nonindigenous species (Wilcove et al. 1998, Stein et al. 2000). Invaders can also interact with habitat transformation and thus exacerbate the threat to biodiversity (Hobbs 2000).
Extinction of native species, although dramatic, actually characterizes relatively few invasions (Simberloff 1981). Reduced population sizes and local extirpation of a species appear more common than global extinction of a species, but changes in population sizes of native species after invasion by nonindigenous species can vary greatly in magnitude and even direction. For example, establish-
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
native species richness that the original native community no longer exists and has been supplanted by a new community dominated by one or more invasive species. Even the emergence of the new community does not guarantee the end of alteration. In the western United States the once dominant role of Bromus tectorum is being reduced locally through the more recent invasion of Centaurea solstitialis. Thus, one invader is being supplanted by another (R.N. Mack, personal observation).
In native forest ecosystems, insects and microorganisms are intrinsically involved in such processes as decomposition and nutrient cycling, maintenance of forest productivity, pollination, and food webs (Mattson and Addy 1975, Haack and Byler 1993, Gilbert and Hubbell 1996). Their influence on forest composition and structure occurs on genetic to landscape scales, but these organisms have evolved with their ecosystems over long biological periods. In contrast, introduced insects and microorganisms have no evolutionary history with the forest ecosystems that they have come to influence. Even though most foreign invaders are similar to indigenous organisms in how they feed or infect their hosts, the ecological changes that have resulted have been dramatic and cascading and have occurred over short periods (Gibbs and Wainhouse 1986, Oak 1998).
For example, forests since the invasion of the chestnut blight have often been dominated by oak species. Their increased abundance and continuity now provide ideal hosts for the introduced gypsy moth; for the increased expression of oak wilt, an indigenous vascular disease; and for drought-related declines (Liebhold et al. 1995, Simberloff 1996). Likewise, beech bark disease, caused by a nonindigenous scale insect and a pathogen, has led forests once composed of beech thickets or sites to be transformed to grass or shrub land–a change that can alter fire regimes (Oak 1998). Both maladies have affected wildlife populations that depend on beech nuts (Martin et al. 1961). A nonindigenous foliage feeder such as the gypsy moth may function like a native defoliator and appear to have minor effects on the forest, but the consequences of invasion include cumulative stresses on the host and alteration of the populations of other native herbivore species, and these effects can extend to other trophic levels.
What might be the ultimate, global result of mixing of the world’s biota? Brown (1995) has estimated, on the basis of species area relationships for continents, the worst-case scenario for the impact of free exchange of biotic material across former biogeographic barriers. The estimate assumed that continental drift could be reversed and that all the earth’s land could be reassembled in a single giant continent, but with the current climates and geological features intact. With these assumptions, there would be a massive decrease in species—65.7% in land mammals, 47.6% in land birds, 35% in butterflies, and 70.5% in angiosperms. Brown tempered that pessimistic assessment with a number of caveats. Colonization and expansion of ranges must ultimately decrease global diversity, but they tend in the short term to increase local diversity. The extent of the biotic enrichment varies, but many countries have 20% or more nonindigenous species
in their floras (Vitousek et al. 1996). Deliberately introduced species can play a role in the maintenance and management of ecosystem processes. Examples of such species are natural enemies of pests for biological control; aesthetically pleasing, fast-growing, pollution-resistant horticultural plants; fish communities in reservoirs; and grasses that can reclaim strip-mined land in arid regions. The danger arises from nonindigenous species that either play no constructive role or play unexpected roles in their new ranges.
Invaders that affect ecosystem processes—such as productivity, nutrient cycling, or disturbance regimes—have been viewed as the most difficult to quantify and verify (Vitousek and Walker 1989, Mack and D’Antonio 1998). In a sense, changing ecosystem processes “changes the rules of the game” in a way that influences many, if not all, of the component species. Plant invasions can also alter nutrient-cycling patterns, as illustrated by the invasion of the nitrogen-fixing tree Myrica faya on volcanic surfaces in Hawaii (Vitousek and Walker 1989). The invasion of American rangelands by Bromus tectorum (cheatgrass) has increased the frequency and intensity of fires, thereby transforming steppe once dominated by the shrub Artemisia tridentata (big sagebrush) to annual grasslands (Whisenant 1990). Similarly, the invasion of nonindigenous annual grasses into Californian chaparral has resulted in more-frequent and more-intense fires, which in turn have altered species composition (Zedler et al. 1983). Plant invasions can also alter hydrology, as illustrated by Melaleuca (Melaleuca quinquenervia), which increases soil elevations and thereby has influenced the hydrology of Florida wetlands (Schmitz et al. 1997), and by the invasion of Pinus spp. into the South African fynbos, which has radically reduced the water yield of catchments (Le Maitre et al. 1995). A recent review (Parker et al. 1999) indicates that most studies of the impacts of invaders on ecosystem processes have concentrated on the effects of the plants—through uptake of light, nutrients, or water—on other plant species. Native animals are also affected by plant invaders (Braithwaite et al. 1989), through loss of habitat and loss of food resources; these interactions have been little studied and might well be underestimated.
An invader would have substantial social or economic effects if it altered “ecosystem services” (cf. Ehrlich and Mooney 1983), such as maintaining the gaseous composition of the atmosphere, controlling regional climates, generating and maintaining soils, controlling floods, disposing of wastes, recycling nutrients, and controlling pests (Ehrlich and Wilson 1991). A potentially global change is under way through the conversion of much of the forested Amazon drainage to grasslands. Huge swaths of tropical forest continue to be cleared, burned, and sown with nonindigenous grasses. These grasses, such as Melinis minutiflora and Brachiaria spp., which were introduced primarily from Africa, are forming a variety of new communities: some appear to require continual
cultivation, and others escape cultivation and are forming invasions. In either case, aggregate effects of the grasslands have enormous impact on the composition of greenhouse gases in the atmosphere and alter the light regime (reflectivity and energy balance) and hydrology in their region and beyond (Mack et al. 2000, Williams and Baruch 2000 and references therein). Perhaps nowhere else on the earth are invasive species altering the biosphere to the same extent, but similar regional transformations are simultaneously occurring elsewhere. The spreads of the invasive Pennisetum ciliare (buffelgrass) in Mexico (Burquez and Quintana 1994 as cited in D’Antonio 2000), of Imperata cylindrica (alang alang) in the tropics and subtropics (Lippincott 2000), and of Pennisetum spp. in Madagascar (P. Binggeli, personal communication) are examples of the replacement of forest or parkland communities with invasive grasses.
The value of lost ecosystem goods and services is often not recognized, because they are not traded on financial markets. But these commodities and services, which are assumed to be available free for all, are under threat. Some progress is being made in calculating the value of ecosystem goods and services and using this information in environmental decision-making and economic policy (Daily et al. 2000). The connections between biodiversity, ecosystem services, and human health must be better understood before a predictive theory of invasion impacts can be developed.
Cumulative and Indirect Effects
The adverse effects of a single invasive species can be small, but the aggregate effects of multiple invasive species can be large. Indirect effects occur when one species influences another via intermediate species, as when two species interact via a shared natural enemy or a shared resource. Interactions among gypsy moths, mice, and Lyme disease in eastern North American oak forests (Elkinton et al. 1996, Jones et al. 1998) illustrate the cascading direct and indirect effects that can occur when communities are tightly linked. In central Massachusetts, oak trees, the preferred host of gypsy moth larvae, produce large acorn crops every 2-5 years. Acorns are an important winter food source for mice, and mice density increases after heavy acorn years. Mice and deer are the primary hosts of the black-legged tick, Ixodes scapularis, which is the vector of the spirochete bacteria, Borrelia burgdorferi, that causes Lyme disease in humans. Heavy defoliation during gypsy moth outbreaks reduces the vigor of oaks, and that results in decreased acorn production, which in turn leads to lower density of mice, which are important predators of gypsy moth pupae in New England. Researchers found that increased abundance of acorns was associated with lower gypsy moth survival but higher densities of mice and host-seeking deer ticks, which presumably increase the incidence of Lyme disease. Such chain reactions are difficult to identify, let alone predict or manage. Moreover, these interactions can vary spatially or temporally. In West Virginia, where gypsy moth popula-
tions have only recently become established, there was no association between abundances of small mammalian predators and gypsy moth dynamics (Grusheky et al. 1998).
The composition and structure of North American forests have always been in flux as they are continuously affected by biotic and abiotic agents. Human activities associated with settlement over the last 400 years have been some of the most pervasive forces in shaping the forests that exist today (Franklin et al. 1987, Merrill 1996). Among those activities have been logging and agricultural practices and the fires that were used to aid hunting and the preparation of land for farming. As the primeval forests were felled, their diversity, composition, and complexity were greatly altered. The forests that have regenerated represent the cumulative effects of the various disturbance agents and are the forests into which most invasive organisms have been introduced (Cronon 1983).
Insects and pathogens historically were viewed as two of the most important damaging agents with respect to forests (Hepting and Jemison 1958). Attempts to eradicate or suppress their outbreaks were initiated when pests produced effects that were in conflict with wood and fiber production, destroyed wildlife habitat, or interfered with natural resources that have been aesthetically valued by humans. More contemporary views have recognized insects and pathogens as normal forest ecosystem components. Their roles as recyclers of carbon and other nutrients, as pollinators and plant symbionts, as food sources for vertebrates, invertebrates, and other microorganisms, and as creators of habitat for wildlife are well summarized (Haack and Byler 1993, Gilbert and Hubbell 1996). In contrast, introduced insects and pathogens are not normal components of the ecosystems that they have come to influence. The general view is that invasive insects and microorganisms are not regulated by the co-evolved resistance mechanisms in their hosts or by the parasites, predators, and diseases that regulate them in their native ecosystems (von Broembsen 1989). All too often, the impact of an invading species has been viewed narrowly as causing extensive mortality and growth loss in the affected species. In reality, the ecological changes that have resulted from their damage typically set off a cascading chain of events that has resulted in rapid ecosystem changes.
A brief examination is warranted of how chestnut blight and gypsy moths have created cascading events to alter forest organization rapidly. Chestnut blight resulted in the most profound set of changes ever recorded in a North American forest ecosystem. The causal fungus, Cryphonectria parasitica is native to East Asia; after its early-1900s discovery in North America, it proceeded to infect and kill American chestnut trees, which once made up 25% of the eastern hardwood forest (Liebhold et al. 1995). As the chestnuts died, the newly available space was occupied by middle-story and understory species, including oak. Fire disturbances before the blight had given oaks an early advantage over light-seeded, less-fire-tolerant, thinner-barked species such as maple and yellow poplar (Stephenson 1986, Oak 1998). As the forests with emergent oaks have
matured, they have been subject to oak decline, a condition that appears to develop as physiologically mature trees experience insect defoliation or drought (Houston 1987). Although the native insect defoliators, such as elm spanworm (Ennomos subsignarius), have been implicated in this decline, the introduced gypsy moth (Lymantria dispar) has been the most important agent of damage. Gypsy moth defoliation disturbs the carbohydrate physiology of oak root systems and makes them highly susceptible to native root-invading fungi (Armillaria spp.) and insects, particularly the two-lined chestnut borer (Agrilus bilineatus), a root collar insect (Houston 1987, Oak 1998). As the gypsy moth has continued its spread southward and westward from New England through the Appalachians, the oak decline-gypsy moth situation has resulted in significant oak mortality. Species that have replaced oaks include more-shade-tolerant trees, such as black gum, red maple, white ash, and yellow poplar. Oak mortality has created dens for wildlife and increased the amount of coarse woody debris. The forests that eventually emerge from the impact of chestnut blight and gypsy moth defoliation may be more tolerant to oak decline and defoliation but will be structurally and compositionally very different plant and animal habitats.
Two other nonindigenous organisms, a scale insect (Cryptococcus fagisuga) and a fungus (Nectria coccinea var. faginata), are influencing the eastern forests of North America as they operate in concert to cause beech bark disease (Houston 1994). The small scale insect creates tiny feeding wounds in the thin bark of beech, which are colonized by the fungus. Eventually, mature beech trees die as the many fungal infections coalesce and girdle the tree. The killing front of this insect–pathogen complex has spread from the introduction point of the scale insect in Nova Scotia to central Pennsylvania; small outbreaks of the disease now occur as far south as the Great Smoky Mountains National Park. The dynamics of beech bark disease begin as infestations by the scale insect and fungal infections of the bark occur along the killing front. As native beeches dies, the remaining beech trees are riddled with nonlethal infections that grotesquely deform them. Further deaths lead to the development of beech thickets that arise from root sprouts. General structural changes in the forest include loss of beech in the canopy, increased snags and downed woody debris, and overabundance of infected beech stems. Long-term effects of the elimination of beech are uncertain, but some stands are already being replaced by sugar maple and yellow birch or have been transformed into grasslands or shrub lands (Oak 1998). Although American beech is not highly prized economically, it had a valuable wildlife role in producing beechnuts, an important food for some birds, squirrels and chipmunks (Martin et al. 1961).
Evaluating Impacts on Communities
There is a fundamental need to identify common standard measures of impacts that would create a more reliable basis for comparison, interpolation, and
extrapolation. What does a reduction by 20, 50, or even 75% in the fitness of a native species mean to its role and persistence on the site, and can these metrics be made meaningful across vastly different taxonomic groups? Similar questions have yet to be addressed about invader-caused alterations in nutrients, water availability, and other ecosystem components. The development of standard approaches to evaluating the impacts of invaders will strengthen future predictive efforts.
The chief reason for the difficulty in evaluating the impact of invasions, however, is the lack of sufficiently detailed data on the species composition, structure, and function of ecosystems before they are invaded. That is clearly the case with the introduction of the chestnut blight fungus: the only information that exists about preblight forests is anecdotal or comes from postblight studies (Stephenson 1986). There can be a long period between an introduction and the spread of an invader (Kowarik 1995). As a result, we recognize that an invasion has occurred only after the ecosystem has changed. The invasion of Bromus tectorum in the western United States, for example, occurred long before the identification of the native plant colonizers that B. tectorum supplanted (Mack 1988). The invasion process is a moving picture, but often we must rely on snapshots that relate spatial variation in structure and function of ecosystems to the abundance of invaders if we are to infer which ecosystem services could be lost as the invasion progresses. Increased availability of data on the composition of natural ecosystems at all levels of complexity would help substantially.
Both the invader and the recipient ecosystem are likely to change. Some invaded systems settle to a new equilibrium; others undergo sustained “boom and bust” oscillations. Some invaders impose constant effects; the effects of others are more dynamic. For example, planned introductions of “predators” for control of arthropods and weeds show a wide range of dynamic behaviors (Hassell 1978): some predators have little detectable effect on their prey’s dynamics; in other interactions, the predators clearly are maintaining their prey at very low equilibrium densities; a few predators and their prey undergo cyclic oscillations; and some interactions are characterized by episodic prey outbreaks when predation ceases to be limiting.
A general increase in the temporal and spatial scale of invasion studies is needed to quantify explicitly cases among each of these categories. For example, large-scale studies are needed to incorporate background variations in response variables and to incorporate variations in changes associated with invasion that are not captured by small-scale studies. One approach is to construct matrices that describe the variances for each species in an ecosystem along the diagonal elements and the covariances between the population sizes of all pairs of species across the off-diagonal elements—at least one matrix for the variances and covariances before the invasion and one for after it (Parker et al. 1999). Comparisons of the changes in such matrices could provide a description of changes in the structure and dynamics of an invaded community.
BOX 5-1 Impacts of Invasive-Species Management Efforts
In the rush to control invasive species, we sometimes create long-term problems that are harder to address. Importation of nonindigenous predators, parasitoids, or pathogens to control invasive pests can have long-term deleterious effects on native, nontarget species. Compsilura concinnata was intentionally established for gypsy moth control in North America. This parasitic fly is known to parasitize over 180 native lepidopteran species; it might be responsible for dramatic declines of large attractive species, such as the cecropia moth (Boettner et al. 2000), and it has expanded well beyond the range occupied by gypsy moths.
Similar examples of unanticipated effects on nontarget species have been documented for insects introduced for biological control of undesirable plants. The European weevil Rhinocyllus conicus, introduced in 1968 to control weedy thistles, now affects native nontarget plants, including rare species of Californian Cirsium (Ehler 1991, Louda et al. 1997). An Argentine moth, Cactoblastis cactorum, was introduced into the Caribbean in 1957 to control undesirable Opuntia species. Its unanticipated arrival in Florida in 1989, however, has generated concern about its impacts on five native Opuntia species, including a rare, protected cactus. Continued range expansion threatens the diverse guild of Opuntia in Mexico (Johnson and Stiling 1996, Pemberton 1995, Simberloff 1992, Strong and Pemberton 2000). Two immigrant ladybird beetles, Coccinella septempunctata (L) and Harmonia axyridis (pallas), originally imported for biological control, are suspected of displacing native ladybird beetle species (Colunga-Garcia and Gage 1998), and H. axyridis is now considered an annoying pest because it aggregates in masses on homes.
A further complication arises when an invasive species is deemed a curse by one segment of society and a salvation by another. That possibility is well illus
We are unlikely to obtain such quantitative information on patterns of variation and correlation among abundances of species in a community, so it is reassuring that simple qualitative analysis and modeling can further our understanding of potential ecological effects of nonindigenous species and serve as a practical tool for adaptive management (Li et al. 1999). A community matrix can be constructed by representing relationships between the species (or variables) in a community with scores of positive (+1), negative (-1), or zero values, signifying the effect of one species or variable on another. Mathematical tools can be applied to determine the presence of equilibrium and its stability, as well as the effect of input into the system. This approach has led to important policy and management recommendations in agriculture, and fisheries, and has been useful in risk analysis of species introductions.
trated by the protracted legal battles in Australia over Echium plantagineium, given the antithetical sobriquets of “Patterson’s Curse” and “Salvation Jane” (Parsons and Cuthbertson 1992)! A similar situation may yet arise in the United States over the herbaceous species Hypericum perforatum (St. John’s wort). It is unquestionably an aggressive rangeland weed and was the target of a successful biological control program (Huffaker and Kennett 1959). But today the plant is valued by some as an herbal remedy and is touted as a boon to some local economies.
Furthermore, the “zero-tolerance” policies used by many regulatory agencies might require pesticide applications, inspections, or other treatment before plant material from areas known to be infested by a nonindigenous insect or pathogen is permitted to enter uninfested areas. Such policies are incompatible with biological control, nullify economic injury or action thresholds that are integral components of integrated pest-management programs, often lead to increased pesticide use, and can result in considerable costs to producers, inspection agencies, and other officials in affected areas.
Even attempts to disrupt the life cycle of the nonindigenous pest without classical biological control can have environmental consequences. For example, establishment of white pine blister rust had profound ecological impacts in many northern states, but the effects of programs to eradicate the native Ribes species (the alternative host of the pathogen) and discourage regeneration of white pine have not been addressed.
Collectively, those examples show the need for careful evaluation of biological control agents, inasmuch as protocols for screening the evolutionary potential of invaders–weighing genetic variation, natural selection, and ecological opportunity—are still in early stages of development (Ewel et al. 1999). The potential for such unintended consequences and the public’s diverse reaction to nonindigenous species need to be assessed as much as possible before the species’ release.
Of the country’s persistent nonindigenous species, only about 10 % are considered invasive. But appearances can deceive, and many of the remaining 90% might be considered innocuous only because their harmful effects have not been documented or even investigated.
Effects or impacts of invasive species are often hard to measure and even harder to predict, because scientific uncertainty arises in quantifying each step in the invasion process. Moreover, data are lacking on species composition and species abundance in many ecosystems before the arrival of invasive species.
Impacts occur on various scales—biological, spatial, temporal. Effects become harder to predict as one moves from the individual to the genetic, popu-
lation, community and ecosystem levels. Long-term evolutionary and community effects are the least well studied and perhaps the least predictable. Furthermore, there are few widely accepted measures of the impact of invasive species.
Classification of organisms by functional groups might help in predicting impacts. Among invaders, grasses that produce massive amounts of combustible fuel, plants that form mutualisms with nitrogen-fixing bacteria, and trees that produce dense, light-diminishing shade can have huge impacts in ecosystems in which these functional groups had been absent or their effects have been negligible.
Invasive species often exert their influence by initiating a cascade of changes in the biotic and abiotic components of the ecosystem; examples are chestnut blight and gypsy moths in eastern North America.
Some plant invasions, such as in the Amazon watershed, are becoming so extensive that they are probably affecting global atmospheric circulation and the global carbon budget.
Few studies have been conducted on large temporal and spatial scales. Such studies could incorporate background variations in invasion-associated changes that are not captured by small-scale studies.