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More than 20 species of exotic plant pathogens have been introduced into forests in the United States (Pimentel and others 1999) and exotic forest pests have greatly altered the species composition of forests in the East (Campbell 1997). Some tree species important as habitat for forest wildlife have been virtually eliminated throughout their ranges or greatly reduced in number. The loss of nuts and berries formerly produced by vanishing or severely reduced tree species has had a poorly documented but surely substantial impact on forest wildlife species (Campbell 1997). See Chapter HLTH-2 for a complete discussion of forest timber pathogens and diseases. Although the impacts of exotic plant pathogens to timber resources are well documented, the impacts on forest wildlife resources are not well described.
At the beginning of the 1900s, the American chestnut was one of the most important wildlife plants of the Eastern United States (Martin and others 1951). With this tree practically exterminated by the exotic chestnut blight, mast-dependent forest wildlife such as white-tailed deer and black bears had to settle for inconsistent acorn and hickory nut crops as their primary food (Clark and Pelton 1999). The blight almost certainly reduced the carrying capacity of southern highland habitats for mast-dependent wildlife. The blight is thought to have caused at least five indigenous insect species to become extinct or extremely rare (U.S. Congress, Office of Technology Assessment 1993). In areas where resprouting chestnuts remain in the understory, birds and mammals continue to transport virulent and hypovirulent-like strains of chestnut blight fungus (Scharf and DePalma 1981). Chinquapins in southern forests (including the Allegheny and Ozark chinquapins) vary in their susceptibility to chestnut blight. The chinquapins may not match the former value of the American chestnut in their habitat contribution to wildlife in southern forests (Martin and others 1951), but the nuts they produce are valuable to wildlife (USDA Forest Service 1999). Chestnut blight has affected chinquapins in southern forests and is expected to continue reducing the prevalence of susceptable tree species. However, no extermination of any southern wildlife species has been documented in conjunction with chinquapin losses.
Dutch elm disease devastated American elms as it spread across most of the country. In areas where Dutch elm disease removed the elm trees from the forest canopy, bird population surveys documented high local extirpation and colonization rates by bird species during the early 1950s (Whitcomb and others 1981). In Great Britain, reductions in bird abundance and diversity were documented in wooded farmlands accompanying elm death from Dutch elm disease and subsequent felling of dead trees (Osborne 1982, 1983, and 1985). The combination of Dutch elm disease and logging reduced the availability of suitable nesting cavities for cavity-nesting waterfowl species (Johnsen and others 1994).
Other exotic plant pathogens continue to affect wildlife habitat in southern forests by reducing the abundance of valuable forest tree species. These include dogwood anthracnose and butternut canker. Flowering dogwoods are valuable to many wildlife species for their fruit production (Martin and others 1951 and USDA Forest Service 1999). Butternuts are consumed by many species of forest wildlife.
Some troublesome weed pests (such as Johnson grass, multiflora rose, and kudzu) were intentionally introduced as crops, for wildlife enhancement or for erosion control but later became pests (Pimentel and others 1999). The majority of weeds, however, were accidentally introduced with crop seeds, from ship-ballast soil, or from various imported plant materials, such as ornamental plants. Some exotic invasive plants such as Chinese privet are shade tolerant and once established are capable of invading relatively dense forests. Many other invasives such as kudzu, mimosa tree, or princess tree are less adept at colonizing deeply shaded, mature forests except along edges, in natural or manmade forest canopy openings, or in disturbed or fragmented forests. Exotic plants have been spread by overgrazing, land-use changes, application of fertilizers and the use of agricultural chemicals (Westbrooks, 1998). Other human activities result in disturbed environments and encourage invasive plants. These activities include farming, creation of highway and utility rights-of-way, clearing land for homes and recreation areas such as golf courses, and constructing ponds, reservoirs, and lakes.
Millions of acres of forest land in the Southeast are occupied by exotic invasive plants. For many species, the acreage infested and spread rates are unknown. Kudzu and Japanese honeysuckle occupy more than 7 million acres each and their spread rates are increasing (Miller 1997). Clearcuts in the South can become infested with exotic vines such as Japanese honeysuckle and mile-a-minute, which can prevent the growth of seedlings and retard timber yields (Campbell 1997 and Nuzzo 1997). English ivy and Japanese honeysuckle can overgrow and eventually kill trees and under story plants and have fundamentally altered the character and structure of some forests (U.S. Congress, Office of Technology Assessment 1993). The herbaceous or shrub layers of large (but undetermined) areas of forest are being transformed into virtual monocultures by exotic vines, herbs, and shrubs (Campbell 1997). In some cases, these plant invasions have been shown to reduce forage or cover for wildlife. Table 1 lists some exotic plant species that are particularly noxious in forests in the Southern United States.
In recent years the impact of invasive exotics on biodiversity has become a major concern. Biological invasions by exotic species may displace native animals and plants, disrupt nutrient and fire cycles, and change the patterns of plant succession (Westbrooks 1998). Invasive exotic plants encroach into parks, preserves, wildlife refuges, and urban areas. Since many of these areas are significant for maintaining indigenous animals and plants (U. S. Congress, Office of Technology Assessment 1993), the responsible land management agencies are forced to expend increasing resources to control the most troublesome invaders. Approximately 61 percent of our National Parks have at least a moderate level of exotic plant infestation: severely impacted parks include the Great Smoky Mountains. An estimated 400 of 1,500 vascular plant species in the Great Smoky Mountains National Park are exotic and 10 of these are currently displacing and threatening other species in the park (Pimentel and others 1999). Invasive exotic species are considered to be the second most important threat to biodiversity, after habitat loss and degradation. Approximately 42 percent, or about 400, of the 958 species that are listed in the United States as threatened or endangered under the Endangered Species Act are at risk because of competition with or predation by exotic species (Wilcove and others 1998). In south Florida, exotic plant species such as Australian pine, Brazilian pepper, and leatherleaf fern are invading disturbed areas and outcompeting native vegetation, reducing Key deer foods and habitat (U. S. Fish and Wildlife Service 1999). In spite of the severity of exotic plant invasion in southern forests, the impacts to forest wildlife in the South have only been sparsely documented. More information about the effects of exotic invasive plants on forest ecosystems can be found in Chapter TERRA-2.
Many exotic invasive plant species lack insect herbivores adapted to live and feed on them. This factor likely contributes to their rapid spread. The number of plant-feeding insects associated with various trees is a reflection of the cumulative abundance of that tree throughout geological history (Southwood 1961). Recently introduced exotic tree species generally support relatively few insect species compared to abundant native tree species. The Chinese tallow tree is an invasive exotic that has spread rapidly across the Southern United States. Insects likely control the spread of this tree in its native China, and the lack of insect predation has aided its spread in the United States. Only one species, the leaf-footed bug, has been reported causing fruit damage to this exotic tree (Johnson and Allain 1998).
Despite the tendency of some exotic plant invaders to form dense monocultures that exclude native flora and fauna, many species of southern wildlife use exotic plant species for forage and cover. Indeed, some invasive plant species in southern forests were introduced because they were considered beneficial for wildlife habitat (Miller 1997).
The value of Japanese honeysuckle both as cover and a food source for songbirds, gamebirds, hummingbirds, small mammals, and deer has been documented (Martin and others 1951, Hugo 1989, Miller 1997). Other exotic honeysuckles such as Amur honeysuckle also have been documented as food and cover for birds and small mammals (Whelan and Dilger 1992, Williams and others 1992, Martin and others 1951).
Multiflora rose is an invasive exotic shrub that was widely promoted by conservation agencies in the 1930s for cover, wildlife food, and as living fences (Miller 1997). It provides excellent habitat for gamebirds and songbirds (Martin and others 1951, Morgan and Gates 1982) and for cottontail rabbits (Morgan and Gates 1983).
Japanese and Chinese privets are invasive exotic shrubs that can replace native understory species and prevent forest regeneration in riparian forests and bottomland hardwood-pine forests (Miller 1997). Privets are used for food and habitat by birds and their seeds are widely dispersed by birds (Martin and others 1951, Miller 1997). Chinese privet also has been documented in northwestern Georgia as an important component of fall and winter diets of the white-tailed deer (Stromayer and others 1998).
Exotic shrubs in the buckthorn family provide excellent nesting and feeding habitat for many species of songbirds (Whelan and Dilgar 1992). The exotic shrub bicolor lespedeza provides food for songbirds, gamebirds, and hooved browsers, including white-tailed deer (Martin and others 1951, and Miller 1997).
The Chinese tallow tree in coastal South Carolina is used heavily by more than 14 bird species (Renne and others 2000). The Russian olive provides feeding habitat for songbirds, gamebirds, and hooved browsers (Martin and others 1951). Chinaberry is eaten to a limited extent by songbirds (Martin and others 1951).
Although these exotic invasive plant species provide habitat and food for southern wildlife species, no scientific investigations were found that compared the relative habitat value of these exotic invaders to the native flora that they displaced. In addition, no scientific investigations were found that documented the effects of exotic plant species invasions on a broad spectrum of southern forest wildlife species, including sensitive habitat specialists. The past introduction of exotic plants for wildlife management has unintentionally led to severe invasive exotic species problems. Many of the intended habitat benefits of these invasive species can be found in carefully selected native species. See the National Park website at http://nps.gov/plants/alien/fact.htmfor some suggested native plant alternatives. Introduction of exotic plant species for wildlife enhancement should be approached with caution to avoid future invasive species problems.
More than 2,000 arthropod species and 11 earthworm species have been introduced into the Continental United States, including approximately 500 exotic insect and mite species (Pimentel and others 1999). About 360 exotic insect species have become established in American forests and approximately 30 percent of these species have become serious pests. Although the negative effects of invertebrate pest species such as the gypsy moth and the balsam woolly adelgid to southern forests have been well documented (see Chapter HLTH-2), much less information is available about their effects on wildlife. See Chapter HLTH-2 for a description of the effects of insects and other forest pests on southern forests.
Balsam woolly adelgid
The balsam woolly adelgid is an aphid that inflicts severe damage in balsam-fir forests (Pimentel and others 1999). The balsam wooly adelgid has killed up to 95 percent of the Fraser firs in the Southern Appalachians.
Resultant habitat losses have impacted forest wildlife. A few species, such as the larvae of the moth Semiothisa fraserata may depend exclusively on the Fraser fir for food (Stein and Flack 1996). Other species such as the Weller's salamander are endemic to the spruce-Fraser fir habitat of the Southern Appalachians. Changes in the avifaunal composition of Fraser fir forests were documented in the Southern Appalachians following destruction of the Fraser fir canopy by the balsam wooly adelgid (Alsop and Laughlin 1991, Rabenold and others 1998).
Frazier fir bark provides substrate for eight rare species of mosses and liverworts (Stein and Flack 1996). The endangered spruce-fir moss spider lives in moss mats that are only found in the spruce-Fraser fir forests of Southern Appalachia (U.S. Fish and Wildlife Service 1998). Loss of the tree canopy (due to the balsam woolly adelgid) has resulted in increased light and temperature and decreased moisture on the forest floor, causing the moss mats on which the spider depends to dry up and become unsuitable.
The endangered Virginia northern flying squirrel and the endangered Carolina northern flying squirrel are found in conifer-hardwood ecotones or forest mosaics of spruce-fir associated with various hardwoods in high elevations of the Southern Appalachians (U.S. Fish and Wildlife Service 1990a). Although decimated by past logging of spruce forests, these two subspecies are currently threatened by several factors including habitat damage to conifer-hardwood ecotones by the balsam wooly adelgid and gypsy moth.
The gypsy moth was accidentally released in Medford, Massachusetts, in 1869. The spread rate of gypsy moths from 1966 through 1990 was approximately 13 miles per year (Liebhold and others 1995). Gypsy moths feed on numerous trees, shrubs, and vines, but prefer oaks (USDA Forest Service 1999).
Infestation by gypsy moths can impact forest wildlife habitat in several ways. Severe infestations can reduce the production of acorns and mast produced by susceptible tree species, reducing mast available for wildlife. However, resultant dead trees can serve as dens for some wildlife (Brooks and Hall 2000). Defoliation of the overstory can displace closed-canopy bird species while increasing the abundance of open-canopy species (Michigan State University, website for education program 1997). In some heavily overstocked forests lacking natural disturbances (such as fire), defoliation can benefit forest birds dependent upon smaller openings in mature hardwood or mixed forests. Beneficiaries include some declining or priority species such as Canada warblers and white-throated sparrows (Hunter and others 2001).
Following gypsy moth infestations, sensitive shade-dependent understory plants can become stressed by the increased sunlight reaching the forest floor (USDA Forest Service 1999). Defoliation of the overstory increases the growth of shrubs, grasses, and herbs providing some wildlife with additional cover and forage (Brooks and Hall 2000).
Red imported fire ants
The red imported fire ant infests more than 250 million acres in the United States (Allen and others 1994). Fire ants could spread across almost a quarter of the nation before range limits are reached. Southern States already infested by the species suffer damages totaling more than $1 billion per year (Pimentel and others 1999).
Red imported fire ants are most abundant in open habitats with disturbed soil, where sunlight can reach the soil surface (Stiles and Jones 1998). They are rare in shaded or undisturbed habitats such as intact forests. Fire ants can invade southern forests along the margins of linear disturbances such as roads or powerlines. In areas where the red imported fire ant is abundant, native ants are displaced by competition. Although omnivorous, the species feeds voraciously on living and dead insects. Native arthropod diversity and abundance often are reduced in heavily infested areas (Tedders and others 1990, Stiles and Jones 1998, Allen and others 1994).
Red imported fire ants have had detrimental impacts on many wildlife species (Allen and others 1994). Reptiles and amphibians tend to be vulnerable to displacement by fire ants when they compete for shared prey (invertebrates) or have an egg stage vulnerable to predation during times of high fire ant activity. Fire ants have been documented to destroy nests and cause hatchling mortality of the threatened gopher tortoise (U. S. Fish and Wildlife Service 1990b, Allen and others 1994).
Fire ants compete with native scavengers that feed on dead animals and fallen fruit. They have been implicated in declines of ground-nesting birds, such as quail and turkey, because they attack newly hatched young (USDA Forest Service 1999). Nest and chick predation by the red imported fire ant has been documented for many bird species (Allen and others 1994). The red imported fire ant has been linked to declines of migratory wintering populations of the loggerhead shrike (Grisham 1994). Injuries or death to white-tailed deer fawns and other new-born small mammals due to attack by the red imported fire ant have been widely reported (Allen and others 1994).
Stein and Flack (1996) estimate that at least 2,300 species of exotic animals now inhabit the United States. This total includes an estimated 20 species of exotic mammals, 97 species of exotic birds, and 53 species of exotic reptiles and amphibians. These species cost the U. S. economy about $27.5 billion every year (Pimentel and others 1999 and Scientific American 1999). Many of the larger exotic animals were deliberately imported for aesthetic, sport hunting, or livestock purposes. Deliberate imports include European starlings, European wild boars, ring-necked pheasants, and feral pigs. Other smaller exotic pests, such as rats, mice, red imported fire ants and balsam woolly adelgid arrived hidden in cargo holds, shipping containers, produce, and imported forest products. Echternacht and Harris (1993) indicated that at least 50 exotic wildlife species have become established in the Southeastern United States comprising about 8 percent of the 625 native and exotic wildlife species. Table 2 is based on their wildlife and faunal description. It contains a list of exotic wildlife species that are known to inhabit the Southeastern United States.
Feral pigs that descended from domestic farm animals and European wild boars that were introduced for sport hunting now number about 4 million across the United States. Together, they cost the economy more than $800 million in damages per year (Pimentel and others 1999). Florida has about 0.5 million and Texas has 1 to 1.5 million.
The effects of wild pigs vary greatly from place to place, depending on the density of pigs and the sensitivity of the ecosystems involved (Singer 1981). Their rooting habit has damaged sensitive forest habitats across the South, including rare wetlands and springs in the Ozark-Ouachita Highlands (USDA Forest Service 1999). Wild pigs compete with wild turkeys and white-tailed deer for acorns and other foods. They tear up rotten logs that provide habitat for many amphibians and reptiles. In addition, hogs destroy the nests of turkeys, ruffed grouse and other ground nesting birds (Sealander and Heidt 1990, Miller and Leopold 1992). Wild pigs also carry diseases such as brucellosis and pseudorabies that represent a risk to native wildlife (Peine and Lancia 1990, New and others 1994, Tozzini 1982). No antibodies for serious diseases were detected in a 1990 survey of wild pigs in the Great Smoky Mountain National Park, however (New and others 1994).
Wild pigs occur in 13 National Parks but are especially problematic in the Great Smoky Mountains National Park (Singer 1981). Wild boars invade high-elevation northern hardwood communities from about April through August where their rooting has reduced understory plant cover up to 87 percent. Up to 77 percent of all logs and branches are moved in heavily rooted areas. Red-backed voles and shrews are normally common in pristine stands, but are absent in rooted areas.
Domestic cats, including both pets and free-ranging animals, now number about 100 million in the United States (Coleman and others 1997). The occurrence of cats tends to be concentrated around areas of human habitation. Studies of free-ranging domestic cats indicate that small mammals comprise about 70 percent of their prey, and birds constitute about 20 percent. Nationwide, free-ranging rural cats probably kill more than a billion small mammals and hundreds of millions of birds each year. Free-ranging cats are a serious threat to ground-nesting birds such as turkey and quail (USDA Forest Service 1999, Miller and Leopold 1992) and also attack shrub-nesting songbirds. In Florida, free-ranging cats are contributing to the imperiled status of several Federally listed species including the Lower Keys marsh rabbit, several types of beach mice, and woodrats.
Free-ranging cats can outnumber and compete with native predators, including hawks and weasels (Coleman and others 1997). Cat predation may deplete winter populations of microtine rodents and other prey of red-tailed hawks, marsh hawks and American kestrels (George 1974). Free-ranging cats also can potentially transmit new diseases to forest wildlife, including feline leukemia to cougars (Jessup and others 1993) and feline distemper and feline immunodeficiency virus to the endangered Florida panther (Roelke and others 1993).
Free-ranging and feral domestic dogs are nearly ubiquitous across the United States (Drost and Fellers 2000): many problems are reported in Florida and Texas (Pimentel and others 1999). Free-roaming dogs chase and harass indigenous wildlife (U. S Congress Office of Technology Assessment 1993, Sealander and Heidt 1990) and disturb ground-nesting birds such as quail and wild turkeys by attacking adult birds, and consuming eggs and hatchlings (USDA Forest Service 1999, Miller and Leopold 1992). In southeast Alabama, free-ranging dogs prey upon the threatened gopher tortoise and destroy gopher tortoise burrows (U. S. Fish and Wildlife Service 1990b and Causey and Cude 1978). In south Florida, dog-related deaths are the second most frequent cause of man-induced mortality for the endangered Key deer (U. S. Fish and Wildlife Service 1999)
Free-ranging dogs have the ability to interbreed with coyotes and the Federally endangered red wolf (Sealander and Heidt 1990 and USDA Forest Service 1999).
After the introduction of European starlings in the late 1800s, population growth and range expansion were explosive. Starling populations now appear to have leveled off or are decreasing in most areas across the country (Robbins 2001). Although starlings consume noxious insects and weed seeds, they also compete with native species for food and nesting cavities. Displacement of native birds by starlings has been documented in areas of the country with limited nest sites (Weitzel 1988). Starlings are known to be a very aggressive species when competing for or usurping cavities from other birds (James and Neal 1986).
Effects on reproduction and fecundity of red-bellied woodpeckers were documented due to nest cavity competition with starlings (Ingold 1994, 1996, Ingold and Densmore 1992). The effects of starling nest cavity competition on northern flickers and red-headed woodpeckers were found to be less severe. Competitive cavity losses for red-headed and northern flickers have more serious implications, however, since these two species are currently declining. Starlings are common in urban and agricultural woods, but are seldom found in densely forested areas (Ingold and Densmore 1992). Red-bellied woodpeckers that nest in more heavily wooded environments are more successful in avoiding competition with starlings. Starlings also compete with other native birds, including the eastern bluebird and purple martin for cavity nest sites (USDA Forest Service 1999).
Following a series of introductions in the United States, house sparrows became well established across the continent by 1910. Currently, populations appear to be stable or decreasing in most areas of the country (Robbins 2001). House sparrows are found mainly in urban and agricultural areas (James and Neal 1986) and are seldom found in predominantly forested areas.
Although they commonly nest in man-made structures, house sparrows also use deteriorating nests of other species, woodpecker cavities, and nesting boxes intended for other species. House sparrows have been documented to usurp cavities from red-bellied and red-headed woodpeckers (Ingold and Densmore 1992). In addition to native woodpeckers, house sparrows have been known to harass other native birds including robins, yellow-billed cuckoos, and black-billed cuckoos. They can displace native eastern bluebirds, wrens, purple martins, and cliff swallows from their nesting sites (Arcieri 1992, Pimentel and others 1999). The deaths of adult and nestling bluebirds were documented in South Carolina resulting from aggressive competition with house sparrows (Gowaty 1984).
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