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It is very important to view terrestrial ecosystems at a landscape level. Substantial research has been done on the effects surrounding landscapes have on the health and status of migratory birds, salamanders, and black bears. Below are summaries of our present understanding of the complex relationships for these groups of species.
Since the 1970s, biologists have been documenting the decline of migratory bird species from isolated woodlots and parks nestled in agricultural- or urban-dominated landscapes in the Midwestern and Northeastern United States (Harris 1984, Robbins and others 1989, Robinson 1992, Temple and Cary 1988, Terborgh 1989). These local declines have been attributed to forest fragmentation, where negative effects on populations occur due to increasing isolation of what otherwise should be suitable habitat.
Among the negative effects, the best documented are factors that reduce reproductive success, especially those associated with elevated nest predator and nest parasites like the brown-headed cowbird populations (Brittingham and Temple 1983, Dijak and Thompson 2000, Gates and Gysel 1978, Keyser and others 1998, Rich and others 1994, Robinson 1992, Wilcove 1985). However, for birds that have high dispersal capabilities, it is theoretically possible for "sink" populations—those with reproduction below which a populations can be sustained—to be large and seemingly "stable" (Pulliam 1988). The persistence of some migratory bird populations in the face of reduced reproductive success is usually explained by the immigration of individuals from more secure populations (Robinson 1992). These more secure "source" populations of forest birds, where reproduction supports a surplus of individuals, presumably are from more largely forested landscapes. In theory, the more isolated the sink population from source populations, the more likely that sink population will eventually collapse.
Other factors associated with forest fragmentation may affect birds, but are more important for other wildlife species less able to widely disperse. These other factors include: (1) increased mortality of individuals moving between patches, (2) lower recolonization rates of empty patches, and (3) reduced local population sizes resulting in increased susceptibility of species to regional extirpation or rangewide extinction (Trzcinski and others 1999). Recent studies also have documented reduction of food or other vital factors in forest fragments compared with larger, more intact habitats (Burke and Nol 1998).
Many of the negative effects to birds from forest fragmentation are associated with edges between habitat types. Edges between major habitat types can be extremely productive in terms of diversity of cover and food resources. However, predator and cowbird populations often are elevated in edges. Therefore, nesting birds that are attracted to habitat near edges may be overwhelmed by predators or cowbirds. Gates and Gysel (1978) coined the term "ecological trap" to describe situations where nesting attempts are doomed to failure (also see Donovan and Thompson 2001).
Area-sensitive species do not occur in habitat patches below a certain size. Forest-interior species are usually found in extensive areas of forest interior rather than a diversity of successional stages (Ambuel and Temple 1983, Blake and Karr 1987, Freemark and Collins 1992). However, whether any one species is area-sensitive or associated only with forest interiors varies considerably from place to place, often with respect to the surrounding land use patterns.
Most of the studies cited above were done in the Midwest and Northeast. Relatively few studies in the Southeast have duplicated the long-term studies in other regions, but there is no obvious reason not to apply findings in the Southeast [see Southern Appalachian Assessment (Southern Appalachian Man and the Biosphere 1996) and Ozark-Ouachita Highlands Assessment (U.S. Department of Agriculture Forest Service 1999)]. Results of forest fragmentation studies from landscapes dominated by agriculture and development, however, are not easily transferred to landscapes dominated largely by forest, whether actively or passively managed (Donovan and others 1997; Farnsworth and Simons 1999; Gale and others 1997; Graves 1997; Hagan and others 1996, 1997; Harris and Reed 2001; King and others 1996; Lichstein and others 2002; Marzluff and Ewing 2001; Robinson and others 1995; Simons and others 2000; Wilcove 1988).
Meta-analysis of bird studies across the Midwest suggests that as long as 70 percent forest cover is maintained in largely forested regions, daily nesting survival rates are sufficient to support source populations (Donovan and others 1997, Robinson and others 1995). Where forest cover falls below 70 percent, these and other data suggest that populations may not be sustainable, but large forest patches within a more fragmented landscape may be still able to support healthy populations. Thus, the larger the patch the more species can be supported locally (Robinson 1996).
There is little evidence of negative effects on forest birds in habitats fragmented by various silvicultural methods and associated land uses like temporary roads (Barber and others 2001, Dugay and others 2001, Hartley and Hunter 1998, Villard 1998). There are exceptions involving subtle negative edge effects for otherwise common, stable or increasing, and widespread bird species (Flashpohler and others 2001a, 2001b; Haskell 2000; Manolis and others 2000; Ortega and Capen 1999; Pornezuli and Faaborg 1999; Pornezuli and others 1993; Rosenberg and others 1999), which may reflect subtle changes in habitat condition--more so than habitat fragmentation. On balance forest bird conservation does not have to be focused on fragmentation issues in the Southeast, where overall forest cover exceeds 70 percent in entire physiographic areas (Southern Appalachian Man and the Biosphere 1996, U.S. Department of Agriculture Forest Service 1999).
Therefore, fragmentation is not considered a serious issue for migratory birds in the southern Blue Ridge and northern Cumberland Plateau and Mountains within the Appalachians and much of the Ozark and Ouachita Mountains (Hunter and others 2001). Even in these largely forested areas, local fragmentation due to urbanization may occur, as demonstrated in the southern Blue Ridge and Ozarks (Fitzgerald and others, in press; Holt 2000). Forest fragmentation from agriculture and development is most serious in the Ridge and Valley within the Appalachians, the Piedmont Plateau, the Interior Low Plateaus (outside the western Highland Rim), and the Mississippi Alluvial Plain. Much of the Coastal Plain is intermediate in its percentage of forest land cover, with forest concentrated along the lower Coastal Plain and along major river systems, often including large forest industry tracts.
Pond-breeding salamanders require access from terrestrial habitats to vernal ponds or Carolina bays. Based on a literature review, Semlitsch (1998) recommended for several species of Ambystoma salamanders that buffers around breeding ponds extend to over 160 m (500 feet) and suggested that these areas provide for foraging, growth, maturation, and maintenance. However, even this strategy may not ensure population stability or dispersal among populations unless corridors or connections across the landscape are maintained. Corridors are vital if the surrounding land is hostile to salamander dispersal when timber is removed.
Chazal and Niewiarowski (1998) kept recently metamorphosed mole salamanders in field enclosures. No detrimental effects were detected for animals in recent clearcuts compared to animals in 40-year-old pine stands. These authors hypothesized that the removal of vegetation may not be as detrimental as the mechanical process by which the vegetation is removed. In contrast, Means and others (1996) show that conversion from a relatively open longleaf-wiregrass community, subject to regular burning, to a densely stocked and bedded slash pine plantation can be extremely detrimental for dispersal and access to breeding ponds by the federally threatened flatwoods salamander.
For Plethodontid (woodland) salamanders, there is much conflicting interpretation of data on population responses to clearcutting in montane habitats (Ash 1997, Ash and Pollack 1999, Herbeck and Larsen 1999, Petranka 1999, Petranka and others 1993). Steady return of populations to preharvest levels suggests that fragmentation in largely forested areas is not a serious problem. However, net change in habitat quality may be a serious issue. Important habitat components like substantial coarse downed woody material may be lacking in young stands. Failure of woodland salamanders to reoccupy suitable habitat as it develops or local declines occurring in suitable habitat would be evidence of effects associated with habitat fragmentation, which could lead to population collapse. Thus far, failure of woodland salamanders to reoccupy treated stands remains undocumented, but time lapses may be unacceptably long and the densities reached may be unacceptably low for more vulnerable species.
Fragmentation by roads can seriously restrict movement of amphibian populations. Amphibians on roads die from exposure to predators or are run over by vehicles. Indirect mortality results from lack of suitable habitat facilitating dispersal across roads. Generally, roads of any width and use likely provide some barrier to dispersal. Working in a fragmented landscape, Gibbs (1998) found that most species avoided road-forest edges, but these same species were not inhibited from crossing from forest into fields to reach breeding ponds. In another study by deMaynadier and Hunter (2000), anurans (frogs and toads) were not inhibited from crossing either narrow (5 m) or wide (12 m) roads in a forested landscape; salamanders were inhibited from crossing the wider roads. Thus, in the latter study, wide roads apparently separated salamanders into subpopulations.
Black bears in the Southeast receive a substantial amount of management attention. In addition to a federally listed subspecies in Louisiana, another potentially vulnerable population occurs in Florida. Other healthier populations are subject to hunting that requires careful management attention. Two concerns have been raised about habitat fragmentation for this species: (1) the amount of forested habitat (with a wide range of successional conditions) needed to support a healthy population, and (2) the road density that is too high to sustain a population. In the Coastal Plain, Peninsular Florida, and the Mississippi Alluvial Plain, for example, successful restoration and active management of all the major forested wetland systems would provide significant progress toward what is deemed necessary to secure black bear populations from southeastern North Carolina to Texas.
About 40,000 ha (100,000 acres) of bottomlands, in largely forested condition, are needed to support a population of between 50 and 200 bears, depending on the quality of the habitat (Rudis and Tansey 1995). By the same criteria, a population of about 1,000 black bears would require between 140,000 ha (350,000 acres) and 1,600,000 ha (4,000,000 acres). These areas could include substantial agricultural acreage. Land planted in grain crops is extensively used by black bears as long as escape cover is nearby.
Existing montane population centers such as the southern Blue Ridge in the Appalachians, the Ozarks, and the Ouachitas do not require a minimum acreage to support a healthy population, but bears may avoid heavily used roads or such roads may cause significant mortality (Clark and Pelton 1999). Narrow and infrequently used roads, however, may be heavily used by bears as movement corridors. Road edges that receive direct sunlight may provide substantial amounts of soft mast (fruit) where otherwise closed canopy forests make this important food source rare (Perry and others 2000). Management of narrow or temporary roads (closures and day-lighting) may be more important than the density of such roads in largely forested landscapes.
Fragmentation is a serious problem in shrub-scrub and grassland as well as forest habitat. In fact, many more species are at risk because of fragmentation of shrub-scrub and grassland habitats, rather than with mature forest habitats (Brawn and others 2001, Hunter and others 2001, Larem 1996, Lee and Norden 1996, Litvaitis 1993, Litvaitis and others 1999, Litvaitis and Villafuerte 1996, McCoy and Mushinsky 1999, Opler and Krizek 1984, Woolfendon 1996). These isolated patches of shrub-scrub and grassland habitat may be in agricultural or developed landscapes as well as in forest-dominated landscapes where stocking density has increased (Dunning and others 1995, Means and others 1996).
The challenge for land managers is to improve habitat conditions for a broad array of grassland, shrub-scrub, and mature forest species. Because of differences in land values, this challenge is theoretically more easily met in largely forested areas than in agricultural and developed areas. In heavily fragmented landscapes, attempts to improve habitat conditions for priority species may require segregation of species that depend on mature forests from species that require early successional or scrub-shrub or grassland habitat conditions.
Forest management may contribute to fragmentation of a variety of landscapes, but its effects in forested-dominated landscapes are the most complex. Forest management is designed to influence the composition and structure of forests. Changes in wildlife habitat can be viewed as side effects. As with fragmentation effects, most of the research on habitat relationships in Eastern North America associated with forest management involves migratory birds.
During the latter part of the 20th century, forest cover increased in Eastern North America, while populations of many nearctic-neotropical migratory birds declined. Some researchers speculated that declines were largely attributable to accelerating loss of tropical "wintering" habitats (Robbins and others 1989b, Terborgh 1989). Losses of wintering habitat undoubtedly contributed to declines for a number of species. Recent work suggests, however, that most species of nearctic-neotropical migrants are flexible in use of tropical secondary forest [including especially shade-grown coffee and cacao (chocolate) plantations] and successional habitats (for example, see Krichner and Davis 1992, Sherry 2000).
Another bit of evidence implicating changes in the United States is the substantial variation among southeastern physiographic areas in population trends for many forest species. Among wood-warbler species, declines have been steepest in the heavily forested interior physiographic areas, while populations in the more fragmented and heavily managed lowland physiographic areas have increased (Coastal Plain, Piedmont, Mississippi Alluvial Plain) (James and others 1996). One possible explanation that has not been explored thoroughly is that many forest bird populations may be responding to differences in forest conditions that have developed over the last 30 years (Askins 2001; Hunter and others 2001, in press; Holmes and Sherry 2001; Kilgo and others 1996).
Much of the forest cover increase in the Southeast has been through the expansion of short-rotation pine plantations and the increasing dominance of midsuccessional hardwoods that do not provide high quality habitat for forest migratory birds. (Askins and others 2001, Hunter and others 2001, Trani and others 2001). These phenomena may explain declining population trends in interior physiographic areas. They do not explain the population increases in lowland physiographic areas.
Most of the forest loss in bottomland areas outside the Mississippi Alluvial Plain occurred before the initiation of the Breeding Bird Survey (mid-1960s), so there may have been some response to the return of forests in the Southeast after the 1960s. Substantial losses of forested wetlands in the Mississippi Alluvial Plain during the 1960s and 1970s were attributable to increasing soybean prices. For migratory birds associated with forested wetlands, populations have been stable or increasing while there was a substantial reduction in mature forested wetlands and an increase in younger age classes during the last few decades (Hefner and others 1995, James and others 1996, see chapter 20).
In recent years, close to 100,000 acres of forested wetland in the Southeast have been drained and converted to farmland, pine or hardwood plantations, and industrial and commercial development (Sharitz and Mitsch 1993). In the Southeast, about 45 million acres were once covered by floodplain forests. About 37 million acres remained in 1952, and 33 million acres in 1975. Since then, an additional 2 million acres of forested wetlands were converted to nonwetland uses and another 1 million acres were converted to other wetland types (Hefner and others 1995). Thus, about 30 million acres of forested wetlands remained by 1985. Overall, about 30 percent of the Southeast's historical forested wetlands have been lost. In the Mississippi Alluvial Plain, losses approach 80 percent.
Most of the 70 percent of Southeastern forested wetlands that remain have been cutover at least once, and many are severely fragmented. This fragmentation has further contributed to the decline of many rare but wide-ranging species in the Southeast. Forest-interior and area-sensitive species and those that require large tracts of mature and over-mature wetland forests have been particularly hard hit.
Shrub-scrub (short) and forested (tall) pocosins and Carolina bays support large numbers of bird and amphibian species (Lee 1986, 1987; Moler and Franz 1987). Pocosins and Carolina bays occur in the South Atlantic Coastal Plain of North and South Carolina and Georgia. Originally, pocosin communities in the Southeast covered some 3.5 million acres, about 70 percent are in North Carolina (Richardson and Gibbons 1993). Considerably less than one-third of the original acreage now can be considered intact; another one-third have been irrevocably altered (Richardson and Gibbons 1993). There were probably between 10,000 and 20,000 Carolina bays prior to European colonization, the vast majority in South Carolina. Presently, few Carolina bays can be considered untouched by deleterious human activities. Both pocosins and Carolina bays have been converted to farmland or tree plantations (principally pine) or mined for peat. Areas around Carolina bays are also highly susceptible to commercial and residential development (Richardson and Gibbons 1993).
In the South Atlantic Coastal Plain of North and South Carolina, serious concerns have been raised about conversion of naturally occurring forested woodlands, especially pocosins, to bedded loblolly pine plantations or short rotation forested wetlands. In this case, the presumption was that many species of migratory birds would be significantly harmed by this conversion. However, populations of a majority of these species have been stable or increasing, especially in North Carolina where much of the concern about conversion has been concentrated.
There are many inherent reasons to be concerned about pocosin conversion to pine plantations (Moler and Franz 1987), but migratory birds may be faring relatively well [see “Habitat content (composition and structure) issues: summary assessment of wildlife use of pine plantations” for more discussion]. Among the species that partially or totally contradict expectations are the Acadian flycatcher, red-eyed vireo, northern parula, scarlet tanager, and summer tanager in North Carolina, and the yellow-throated vireo, blue-gray gnatcatcher, yellow-throated warbler, black-and-white warbler, prothonotary warbler, worm-eating warbler, Swainson's warbler, Louisiana waterthrush, ovenbird, American redstart, and Kentucky warbler, in both North and South Carolina (see website on Breeding Bird Survey results for each species, especially refer to trend maps: http://www.mbr-pwrc.usgs.gov/bbs/htm96/trn626/all.html). Only the populations of two species, the wood thrush and hooded warbler, typically associated with mature forest wetlands do not fit this pattern.
Migratory bird use of remaining forested wetlands should be watched closely. Monitoring should focus particularly on swallow-tailed kite, cerulean warbler, and Swainson's warbler, which serve as umbrella species in many forested wetland areas across the South.
In the Mississippi Alluvial Plain, thousands of acres have been reforested in recent years, due to efforts associated with carbon sequestration. When such treatments are contemplated, effects on wildlife should be considered. Pashley and Barrow (1993) provide guidance on managing wildlife in forested wetlands.
Populations of many resident and temperate migratory birds associated with open pine stands are undergoing consistent long-term declines across much of their ranges (Hunter and others 1994, 2001, in press). Many other species of pine associated animals and plants associated with natural stands also are vulnerable. The reason for vulnerability is conversion of natural pine to other forest types and to other land uses.
Harvesting the products of southern pine forests remains a very important part of the southern economy, but the pine forests of today's South are very different from the forests found by European colonists and harvested for naval stores and building materials in the 19th century. Since 1952, extent of natural pine stands in the South has declined from about 70 million acres to less than 35 million acres (chapter 16). The natural pine stands being lost include those dominated by longleaf, pond, and shortleaf pines in the lowland physiographic areas and shortleaf, pitch, and Table Mountain pines in uplands [for the latter see “Habitat content (composition and structure) issues: migratory birds in upland hardwood forests in interior physiographic areas”]. Natural stands of slash, loblolly, and sand pine are also declining, but densely stocked pine plantations are composed mostly of these three species.
The loss of most of the longleaf pine ecosystem has placed many wildlife species at risk in the Southeast (Abrahamson and Harnett 1990, Marion 1993, Stout and Ware and others 1993). At the time of European colonization, longleaf forests covered an estimated 92 million acres stretching from North Carolina to Texas, interrupted only by major floodplain forested wetlands and occasional prairies (Frost 1993, Landers and others 1995). By the 1930s most longleaf pine had been cutover at least once. About two-thirds of former longleaf pine acreage is now occupied by other pine species or has been converted to other land uses (Croker 1987, Walker 1991).
Less than 3 million acres of the original longleaf ecosystem remain. The total is considerably less if systems drastically altered by fire suppression are excluded (see chapter 16). The loss of all but a little of the longleaf pine ecosystem has led to the rarity or endangerment of at least 70 plant taxa, particularly on the Coastal Plain and Florida Peninsula but also on the southern Piedmont and other physiographic areas in the Southeast (Noss and others 1995). Among vertebrate animals, the future of the flatwoods salamander, gopher frog, indigo snake, gopher tortoise, coastal plain fox squirrel, and many other species may well depend on reinstituting growing season fire and restoring the longleaf pine ecosystems.
The loss of fire-maintained shortleaf pine communities is also placing many species at risk (Hedrick and others 1998, Wilson and others 1995). Fire-maintained pond pine stands North Carolina pocosins also places many species at risk (Moler and Franz 1987, Richardson and Gibbons 1993). Sparse stands of sand pine are particularly important component of threatened or endangered Florida scrub communities (Myers 1990). Natural loblolly pine associated with forested wetland communities on bluffs and ridges in floodplains can provide important nest sites for species like swallow-tailed kites and bald eagles. Finally, the loss of fire as a management tool in the Appalachians has led to extirpation of many species and called into question the future of endemic Table Mountain pine communities (Buckner and Turrill 1999, Williams 1998).
Although a large number of species depend on mature southern pine forests, most attention has been focused on one species, the red-cockaded woodpecker. The red-cockaded woodpecker will recover only where large patches of mature pines are managed for the special foraging and nesting habits of this species (U.S. Fish and Wildlife Service 2000). Other species that may be found in shrub-scrub, but optimally use sparsely stocked pine savanna and open pine stands include northern bobwhites, Bachman's sparrows and Henslow's sparrows (winter only). Southeastern American kestrels, red-cockaded woodpeckers, and brown-headed nuthatches may be found if longleaf or slash pines are old enough for cavities.
Cooperating private landowners in the North Carolina sandhills and in areas supporting quail plantations in southwestern Georgia play crucial roles in maintaining relatively healthy (and likely recoverable) red-cockaded woodpecker populations. In these cases, timber production is not necessarily the highest priority land use. Cooperative relationships are also being developed with private landowners who manage mature southern pines for timber production. Such relationships require much care and compromise from all parties. Many stands of mature southern pines (including longleaf) may have been cut and converted to other tree species or land uses earlier than originally planned by landowners who feared government regulations to restore red-cockaded woodpecker populations.
Migratory bird declines in the interior South, especially in largely forested areas, may be due to the way much of the forest cover increase has come about. On public land, management has been largely passive since the massive cutting prior to Federal purchase in the 1930s. Much private land has been repeatedly high-graded, with no or little attention to future stand structure or composition. Both of these approaches to managing forests differ markedly from the intensive short-rotation, even-aged management in the lowland physiographic areas. Unfortunately, passive management and high-grading both have led to a lack of structural diversity in mature forests and a serious lack of early seral habitat for many vulnerable species.
Where a combination of even-aged and uneven-aged regeneration strategies is employed, there is increasing evidence that silviculture conducted in largely forested landscapes provides benefits not only to species requiring early successional stages, but also to a surprising number of species requiring mature forests (Annand and Thompson 1997, Bourque and Villard 2001, Pagen and others 2000, Powell and others 2000, Thompson and others 1992). Several studies have documented the importance of early successional forested habitat for providing food and cover for post-breeding and transient juvenile and adult migratory birds (Anders and others 1998; Kilgo and others 1999; Pagen and others 2000; Perry and others 2000; Suthers and others 2000; Vega Rivera and others 1998, 1999).
Some effects of disturbance frequency on general composition and structure are worth summarizing here. In the South, forests that are the least disturbed by fire and storms are in the protected coves of the Appalachians, principally Cumberland Mountains and southern Blue Ridge. Here, mixed mesophytic forests dominate and the few virgin stands that remain, such as those in the Great Smoky Mountains National Park, match up with expectations of what old-growth forests should look like. Also in the Appalachians, spruce-fir-northern hardwood and hemlock-white pine stands once established have developed over centuries with minimal disturbance. Other relatively undisturbed forests include mixed-mesic forests on the Coastal Plain, such as those on the Apalachicola Bluffs. They also include many types of forested wetlands that are removed from frequent natural floods.
When disturbances occur in today's highly altered forests, the effects differ from what would have been expected prior to European settlement. Presumably, storms of moderate intensity caused gaps in uneven-aged, multi-layered forest stands. Densely stocked stands associated with even-aged or heavily high-graded stands are typically resistant to moderate storm intensity. Extreme storms are likely to cause reinitiation of old-growth stands in a more-or-less even-aged state. They also cause younger even-aged stands to be replaced by new even-aged stands. Autogenic regeneration events are largely missing from today's even-aged or high-graded southern forests. This lack of storm-driven autogenic regeneration in midsuccessional or high-graded forest influences habitats for birds and other wildlife (Hunter and others 2001). A difference between even-aged and high-graded stands is that the former can be converted into more vertically structured stands through prescriptions. In most instances, the only option for diversifying high-graded stands is to first clearcut (i.e., start over) and have in prescription intermediate procedures intended to develop vertical stand diversity over time.
The overall lack of forest structure in many of today's forests may explain why so many bird species respond positively to timber management practices in largely forested areas. Heavy and successful use of clearcuts and forest edges by "forest-interior" or "area-sensitive" species in largely forested regions appears to be a response to the poor structure of extensive forests away from treated areas. Clearly, more research is needed on this topic.
Composition also contributes to habitat quality. Forest composition is constantly changing and should be a primary consideration in largely forested regions in the interior physiographic areas. Serious issues related to composition include: (1) the active conversion of hardwoods to pine; (2) the passive conversion through fire suppression of naturally occurring southern pine stands to hardwoods; (3) the conversion, again due to fire suppression, of oak communities to either mesic hardwoods or white pine; (4) loss of southern Blue Ridge spruce forests; and (5) loss of naturally occurring open habitats such as glades, barrens, balds, bogs and fens.
At one end of the management-intensity spectrum are the passive management strategies now most prevalent on public land. These strategies are causing major changes in forest composition and forest biotic diversity. Passive management is causing abnormally heavy stocking, and fire suppression is causing vulnerable mountain yellow-pine communities (principally Table Mountain and pitch, but also shortleaf and longleaf) to succeed into hardwood communities (Buckner and Turrill 1999, Delcourt and Delcourt 1997). Recent southern pine beetle epidemics have all but eliminated these already vulnerable communities from many areas in the Appalachians. Similarly, oak-hickory stands are being invaded by more mesic hardwood species and white pine. These invasions of more mesic adapted species into more fire-prone conditions may lead to extremely high fuel loads during dry years. In the long run, severe and catastrophic fires will result. Catastrophic fires can further alter forest habitat conditions so that most vulnerable species do not thrive, including disturbance-dependent species in the long-run if these catastrophic events are not soon followed by subsequent prescribed burning to restore appropriate habitat conditions associated with regular fire-return intervals (Delcourt and Delcourt 1997, White and White 1996).
Like other forest types, spruce-fir-northern hardwood forests were harvested near the beginning of the 20th century. The stands that replaced them differ from those prior to harvest. Generally, spruce was replaced by fir from higher elevations and northern hardwoods from below (White 1984). Since a high percentage of the community is in public ownership, it would appear that healthy high-elevation biotic communities can be protected. Fraser fir, however, is threatened by exotic pests, possibly compounded by effects from regional air pollution (Nicholas and others 1999, Rabenold and others 1998, White and others 1993). Some effective restoration probably is possible for red spruce but would require the conversion of existing northern hardwood stands to either spruce or spruce-hardwood mixtures. Some 50,000 acres of such treatment would be needed to reach preharvested forest conditions.
As many as seven forest bird species closely associated with southern spruce-fir-northern hardwood high-peaks forests are effectively isolated from more northerly and western populations. Among these species, the northern saw-whet owl appears to be the most vulnerable to potential habitat loss (Milling and others 1998, Simpson 1992), followed by the black-capped chickadee and the red crossbill. Although widespread elsewhere, the owl and other species restricted to high-peaks forests for breeding in the Southeast need relatively high levels of conservation attention. Northern saw-whet owls respond to nest boxes, which may partially mitigate the loss of high-elevation conifers. Owls also may use other habitat, such as older northern hardwoods and hemlock (Milling and others 1998).
Acreage of pine plantations has increased from 2 million acres in 1952 to 30 million acres today, and an additional 25 million are expected in the foreseeable future. Not surprisingly, the conservation community worries about possible effects on the future sustainability of naturally occurring forests in the South. Although a large percentage of this increase and projected increase comes from retirement of agriculture land (see chapter 16 and chapter 6), there is also a substantial loss of natural pine communities. The loss of natural pine acreage is as much due to fire suppression and clearing for agriculture and urbanization as to conversion to plantation pine. In fact, pine plantations that are invaded by hardwoods often become indistinguishable from natural stands. On many millions of acres, fire suppression since the 1950s has allowed former pine stands to now be classified as pine-oak or even upland hardwood forest types (see chapter 16). So there is no direct correlation between loss of natural pine acreage and increase of plantation pine.
Still, much natural pine acreage and hardwood acreage (both bottomland and upland) have been converted and devoted to efficient growth of short-rotation pine in the South. Although there is general recognition that intensively managed pine plantations are not high-quality wildlife habitats when compared with natural pine and hardwood forests, statements made in several chapters of this report suggest overall that such intensification of management is a positive trend (see chapter 14). Certainly, afforestation of millions of acres of farmland provides for many benefits, from carbon sequestration to water quality improvements. Greater intensity of forest management may allow other forested acres to be set-aside for other purposes, such as wildlife and recreation. However, that intensive forest management actually allows other forest lands to be set-aside or managed for other values, such as wildlife, requires greater scrutiny.
How forests not needed for timber production will be used is unclear at best. Land use trends support that many acres of forest land will be developed, regardless of their productivity. There is no indication that funds would be available to support management of forest lands for wildlife short of commercially viable procedures. Over the last 100 years, many millions of acres of pine and hardwood forests have been left in poor condition for many species of wildlife, including both game and nongame species. Even claims that the present and projected increase in intensively managed pine plantations should bode well for early successional species is highly suspect. High stocking rates (700 to 1000 seedlings per acre), increasing use of fertilizers and herbicides for maximizing pine growth, and reduction of fire as a management tool, among other management changes, essentially have eliminated many of the benefits for early successional species of wildlife that were provided formerly in pine plantations that were less efficiently managed. There certainly is no evidence that steep population declines have been halted or reversed with the expansion of intensively managed pine plantations during the last 30 years. Declining trends continue for important species like northern bobwhite, American woodcock, and many species of high-priority nongame migratory birds associated with early successional habitats (Capel and others 1994, Hunter and others 2001, Krementz and Jackson 1999).
Another major issue in the South is the proliferation of chip mills during the last decade. An important background point is that the chip mills were established in many areas because of poor forest conditions created by repeated past "high-grading"—selective removal of the biggest and best formed trees in hardwood forests. What remains is an unhealthy forest that is poor wildlife habitat. In many of these areas, clearcutting for pulpwood is the first step toward improvement, and chip mills make clearcutting feasible. However, when these hardwood acres are replaced with densely stocked pine plantations, wildlife will not benefit for very long. The alternative often promoted as "environmentally friendly forestry" involves diameter-limit cutting for sawtimber. Diameter-limit cutting, in essence, is a form of high-grading, which was the dominant practice that led to the low-quality hardwood stands found in much of the South.
Management of pine for pulpwood and/or sawtimber need not be as bad for wildlife as is often portrayed. Effects on wildlife involve many factors, including landowner objectives, site quality, and options available for implementing management practices (Melchiors and others, in press). For example, planted loblolly pines in pocosins usually replace stands dominated by pond pine, Atlantic white-cedar, or bays. After pines are established, a manager could provide suitable habitat for many neotropical migrants by retaining a dense hardwood understory and midstory. Reduction in growth and quality of overstory pines would be relatively small.
Notably, nearly all of the forested wetlands lost in coastal North Carolina, much of which was pocosin, were converted to nonwetland uses, including pine plantations (Hefner and others 1995). Although concern for the future of remaining pocosin communities is justified, there is evidence that converting "natural" pocosin vegetation to loblolly pine can have neutral to positive effects on some of the vulnerable neotropical migrants. Neotropical migrant use of these pocosins converted to pine plantation is best when hardwoods are encouraged in the understory and midstory through precommercial and commercial thinnings and infrequent burning (Karriker 1993). Among the species appearing to be stable in these commercial forests are yellow-billed cuckoos, Acadian flycatchers, worm-eating warblers, ovenbirds, and prairie warblers. However, loblolly stands managed for sawtimber under these treatments are still less than 20 years old and have yet to show consistent use by the three highest priority species: black-throated green, Swainson's, and prothonotary warblers. These species require large patches of tall pocosins and other forested wetlands along the South Atlantic Coastal Plain. Optimum management of high-priority, nongame landbirds in pine plantations would include retention of some patches of "natural" pocosin vegetation or otherwise encouraging hardwood understory or midstory development. Conversion from hardwoods to pine or pine-hardwood mix, with appropriate management, is clearly better than no forested habitat at all. For many high-priority neotropical migrants in these habitats, however, restoration and appropriate management of forested wetland conditions would be even better.
The hypothesis that forested wetland species are making the transition to using "bedded" pine plantations is supported by studies in North and South Carolina: (1) in the Parker Tract, Weyerhaeuser Company, NC (Kerriker 1993, Wilson and Watts 1999a); (2) in the Woodbury Tract-Pee Dee River, International Paper Company, SC {Lancia and Gerwin [In press (a)]}; Mitchell and others 1999); and (3) in the ACE Basin, Westvaco Corporation, SC {Lancia and Gerwin [In press (b)]}. The latter two study areas are also the subject of a landscape-level analysis in Mitchell and others (2001). Preliminary results from these studies are promising but long-term benefits depend on maintaining substantial hardwood understories with certain structural characteristics. Heavy bird use of existing woodlands may be temporary as forest management becomes more intensive and hardwood types are replaced by pine. Regardless of the reasons, birds usually associated with hardwood forests are making substantial use of pine plantations in the Coastal Plain of the Carolinas, at least for now.
In the Ouachita Mountains, the USDA Forest Service and Weyerhaeuser Company, among other partners, have embarked on a watershed comparison among passively managed, moderately managed, and intensively managed sites. Preliminary results suggest that large areas under active management likely support a variety of habitat conditions at a variety of spatial scales suitable for many bird species, including many high-priority species associated with both mature forest and early successional conditions (Melchiors and others, in press). The more actively a large area is managed, the more heterogeneous the available habitat, and the less actively managed, the more homogenous the habitat. The latter support surprisingly few mature forest species in numbers higher than those found in more actively managed watersheds (Melchiors and others, in press). In contrast to the Carolina studies, where reproductive rates appear to be consistently high, studies from the Ouachita Mountains and Georgia Piedmont have revealed more complex patterns of nesting success that depend on seral stage, burning regime, and percent canopy versus understory cover (Barber and others 2001, Brunjes 1998, Howell 1998, Raftovich 1998). In addition, heavy and apparently successful use of pine habitats in the Carolinas and possibly elsewhere are generally where sawtimber is the target wood product, where sites have the propensity to support substantial hardwood growth or where maintenance of interspersed hardwood stands are maintained as “ecological legacies.” Data are not available to suggest the same is true for the vast majority of pine plantations, which are managed in very short rotations on very well-drained sites with dense stocking and heavy chemical use.
In conclusion, management options exist in some locations to support healthy migratory bird populations. Study results, however, do not cover the vast majority of pine plantations and how they are managed in the Southeast.
Regardless of whether some hardwood species persist in some pine plantations, priority bird species associated with older pine stands are probably harmed the most by the expansion of pine plantations. Plantation pine stands are too dense, too young, or hardwoods in their understories are too dense for the bird species usually associated with open pine stands that are frequently subjected to prescribed or natural fire. Some of these species may persist in managed pine plantations where hardwood intrusion is controlled and snags are retained (Caine and Marion 1991; Dickson and others 1983; Land and others 1989; Moorman and others 1999; Wilson and Watts 1999a, 1999b,).
For nonavian wildlife, results of studies are also mixed, but similar themes emerge for small mammals and reptiles as found for birds. Working in plantations over former pocosins in eastern North Carolina, Mitchell and others (1995) found that small mammals undergo an initial decline, but later recover to preconversion population levels as long as the plantation emulates, to some degree, the understory structure of the former pocosin. Stand thinning and growing-season burning are essential for maintaining gopher tortoise populations in slash pine plantations in southern Alabama (Aresco and Guyer 1999). Longleaf pines with cavities retained in mature park-like pine plantations in the upper Coastal Plain of South Carolina were used for evening bat roost sites and seemed preferred to potential sites in dense canopied bottomland hardwood, mixed pine-hardwood, or loblolly stands (Menzel and others 2001).
Pine plantations are generally poor wildlife habitat. However, with management adjustments (from less intensive to maintaining natural community types mixed in with plantations) many vulnerable wildlife species can successfully use these commercially driven habitat conditions. At the very least, pine plantations may provide buffers around more natural forested habitats that are clearly better than agriculture or urban areas for hardwood associated songbirds (Kilgo and others 1997, 1998)
Any major change in a forest affects the wildlife that live there. Some changes are caused by purposeful management actions. Others are the result of natural processes (Dickson and others 1993). Managers prescribe treatments to enhance the production of various resources or to promote a forest condition, such as habitat for a particular wildlife species or the quality of a scenic vista.
Different wildlife species and populations react differently to habitat manipulations. Some species are habitat generalists, which have the ability to survive in a wide variety of conditions. Others are habitat specialists, which require specific conditions in order to maintain viable populations. These species have evolved over time to capitalize on unique habitat niches.
An example of a bird habitat specialist is the prothonotary warbler, which needs small cavities in midstory trees or shrubs to successfully nest. Other examples of birds that are habitat specialists include cerulean warblers, Swainson's warblers, and red-cockaded woodpeckers. Habitat generalists, on the other hand, can survive and successfully reproduce in a wide variety of conditions. Examples of habitat generalists include white-tailed deer, raccoons, and coyotes.
Wildlife species also differ widely in mobility. Large vertebrates and birds generally have large home ranges. Black bears have been known to travel over 300 miles, and many birds travel between continents. Many amphibian species, on the other hand, spend their entire lives near the place they were born. Therefore, consequences of changing habitat conditions vary widely among wildlife species.
Timing and energy requirements are extremely important for migratory birds. Favorable weather conditions and adequate food are critical to sustain populations. In the context of forest management, providing as much high-quality habitat as possible is critical. Often, due to localized climatic factors, lands on which migratory species depend are less than optimal. Waterfowl, particularly ducks, are often affected by localized drought, failed seed crops, or extended freezes. When these events take place, it is critical that areas outside of preferred migratory routes provide missing elements. Even though most migratory waterfowl breed in the northern portions of this continent, pair bonding occurs on the wintering grounds. Reproductive success and survival, therefore, depend on the quality and quantity of habitat along the entire flyway, including southern forested wetlands.
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content: Jim Baker and Charles Hunter |
created: 4-OCT-2002 |