In addition to yielding products like timber, wildlife, and fish, which are bought and sold often enough to recognized dollar values, forests have attributes that are highly prized even though people are not accustomed to paying for them. A lack of markets does not make these intrinisic values less important than commodity values. In fact, many would argue that they are more important. This section describes four forest resources with important intrinsic values: old-growth forests, wilderness and roadless areas, biodiversity, and aesthetics. Nonmarket values of the type described here can be classified as either active or passive. The term "active-use value" applies to goods and services used in some activity like recreational fishing, skiing, or camping. The term "passive-use value" includes two categories (Peterson and Sorg 1987, Randall 1992): "existence values," which are things people appreciate without actually using them or even intending to use them (like a distant wilderness or an endangered plant or animal), and "bequest values," which are things people want to remain available for others (such as their descendants) to use and appreciate.

There are many competing definitions of old-growth, but the core concept of old-growth refers to forests that are relatively old and relatively undisturbed. Old-growth forests provide many ecosystem services and human benefits. They are regarded as important in protecting biodiversity and providing habitat for carnivores. Although the southern Appalachians contain one of the largest concentrations of old-growth forests east of the Mississippi River, these forests are restricted mostly to the National Parks and the Wilderness Areas in National Forests. To increase the amount of old-growth clearly requires restoration and management of existing forest. While National Forests may contain the greatest reservoir of forest land suitable for the restoration of old-growth, some proponents of old-growth fear that even this area, if managed as old-growth, is not enough to fully protect biodiversity.

The essence of the old growth concept refers to forests that are relatively old and relatively undisturbed (Tyrell and others 1998).  Although quite a few definitions for old-growth forests have been developed for eastern forests, no wide-spread concensus has been achieved. Why?  Because definitions matter in practice. "Old growth" is as much, if not more, a political concept as it is a scientific concept. How old growth is defined may lead directly to decisions about which forest areas can be commercially harvested and which cannot.  For example, strict definitions of old growth result in smaller acreages off-limits to commercial logging (Leverett 1996).  Conversely, loose definitions may result in larger acreages off-limits to commercial logging.

Most people have an intuitive, common-sense-based concept of old growth.  But it varies depending on personal experiences, cultural background, and educational exposure. Tyrell and others (1998) argue that despite these differences, people do agree on some core characteristics of what old growth is or should be:

"Old, uneven-aged forests, with large trees and lots of dead wood, located on productive sites at low elevations that have not experienced large-scale natural or human disturbance, embody the quintessential old-growth concept" (Tyrell and others 1998).

As one moves away from this core concept of old growth, consensus rapidly disappears due to inherent problems defining old growth. Yet there exists a need for just such a consensus to allow inventorying stands and rating the degree to which they contain old-growth characteristics, prioritizing forests for protection and allocation to the old-growth category, and determining whether and when non-old-growth forests can acquire old-growth status.

Despite this lack of consensus, old-growth definitions still exist (Tyrell and others 1998). There are both conceptual and practical ways to define old-growth. Conceptual definitions create a theoretical framework from which operational definitions and models can be derived. Operational definitions provide clear identification criteria which allow for the classification of individual forest stands as either old growth or not or, more usefully perhaps, provide a metric which estimates the degree of old-growthness a particular stand embodies.

Old growth definitions for several forest types in the southern Appalachians have been synthesized from years of research. Examples of these definitions are available for the follow forest types:

• Mixed Mesophytic Forest
• Oak-Pine, Dry and Dry-Mesic
• BROKEN-LINK BROKEN-LINK Oak, Mesic to Wet-Mesic
• BROKEN-LINK BROKEN-LINK Oak, Dry and Xeric

Old-growth is difficult to define because of (Tyrell and others 1998, Giles 2000):

• Lack of consensus as to which forest types occur in an old-growth condition.
• Difficulty in choosing which attributes to use to characterize old growth.
• The variation that exists in species stature and spacing due to site conditions and elevation.
• Lack of available data about historic old-growth forests.
• Questionable representativeness of remnant old-growth example.
• Changes in forest structure due to altered disturbance regimes since European settlement.
• Influences of pollution, exotics, fragmentation, and extinctions.
• The need to move beyond the stand-level treatment to consider the landscape scale.
• Differences in forest characteristics due to geographic location, e.g., eastern versus western forests.
• Differences in land use history.
• Differences in management objectives.
• Differences in values attributed to old-growth forests.
• Altered disturbance regimes over the last 400 years have changed the bioclimatic background within which historical old-growth forests developed and may inhibit our ability to recreate that exact type of old-growth forest.
• Multiple governmental agencies, non-governmental interest groups, and different, conflicting, and unclear laws and regulations all contending for a definition of old-growth forests favorable for their own agendas.
• The political sensitivity and potential impact of any particular definition of old-growth.

There are a variety of conceptional definitions of old growth. This section summarizes definitions that use structural, functional, historical, utilitarian, and economic attributes to describe old growth.

Structural Definitions of Old Growth

One method of defining old-growth forests uses structural attributes. Some of the structural features of the living component of these definitions include:

• An abundance of large, old trees, recognizable by their asymmetrical shapes, relatively long trunks free of low branches (i.e., in-forest as opposed to open-grown shapes), deeply furrowed or plated bark, signs of heartwood decay, large prominent root structures, flattened crowns with protruding dead limbs, large thick limbs, and trunks often showing a twist that develops with age.
• Majority of late successional tree species that establish in the understory and reach the canopy in tree-fall gaps.
• Wide variation in tree sizes and spacing that reflects years of small natural disturbances.
• Diverse vertical structure that includes multiple layers: overstory trees, understory trees, shrub, herbaceous, and ground layers, each reflecting a broad spectrum of ages.
• Well-developed and diverse herbaceous layer, especially in neutral soils.
• Presence of plant and animal species potentially dependent upon old growth such as an abundance of lichens and fungi in acidic soils.
• Absence of multiple-stemmed trees (Martin 1992).

• Large volume of coarse woody debris (CWD) and standing snags representing a wide variety of and decay classes.
• Tree-fall gaps of various sizes and ages.
• Undulating forest floor with randomly scattered pits and mounds where trees have fallen over and decomposed.
• Undisturbed soils, soil macropores, and thick humus layers.
• Little evidence of disturbance by humans (Leverett 1996, Tyrell and others 1998, Giles 2000).
  

There are several benefits of using structural features to define old growth. Different forests will fit a similar structural definition of old-growth when they might not fit similar functional definitions. Therefore, several attributes of structural definitions are more applicable to more forests than functional definitions (Runkle 1996). Also, many of the structural attributes used to define old growth are the same habitat features required by many threatened wildlife species and therefore are useful for wildlife management (Tyrell and others 1998).

There are also several limitations of structural definitions. The age at which old growth develops and the specific structural attributes that characterize old growth will vary widely according to forest type, climate, site conditions, and disturbance regime (Giles 2000). Therefore, these criteria are not universal; some stands of old growth will possess all of them; others will possess only a few (Leverett 1996). For example, old-growth in fire-dependent forest types may not differ from younger forests in the number of canopy layers or accumulation of down woody material (Giles 2000). In this sense, some forests may fit the functional definition of old growth but may not fit structural definitions.

Functional Definitions of Old Growth

Another method of defining old-growth forests stresses functional aspects of stand development. Using Oliver and Larsons (1996) model of stand development, old growth is the condition achieved following stand initiation, stem exclusion, and understory re-initiation stages when the forest enters the stage of canopy replacement and trees regenerate and grow without the influence of external disturbances (Runkle 1996).

Using this functional approach, two subdivisions of functional old growth are recognized: true old growth and transitional old growth. True old growth occurs when all the trees that invaded immediately after the last major disturbance have died. Transitional old growth occurs where young trees have started growing into the forest canopy but some shade-intolerant species that colonized after the last major disturbance are still alive (Runkle 1996).

Using this functional approach, old-growth is similar to the older term "climax," referring to the mature community that ends a succession (Whittaker 1975). However, because not all forests reach a "climax" state, the usefulness of this definition is restricted to only certain forest types (Runkle 1996).

Historical Definitions of Old Growth

Disturbance history is another attribute used to define old growth. Strict historical definitions of old growth require that no recorded history or discernible signs of direct human disruptions are present in a forest stand. The limited impact of early Native Americans is generally accepted as part of the natural order, with forests disturbed after European settlement excluded from the old-growth definition (Leverett 1996).

Some definitions require that an area has been forested since the beginning of the settlement period. However, this constraint excludes stands that have regenerated from a catastrophic disturbance since settlement, and many ecologists argue that forests that undergo periodic natural disturbances such as windthrow, fire, or insect damage should not be excluded from old growth (Leverett 1996).

Some historical definitions of old-growth relax the prohibition against direct disturbance by humans and include old second-growth forests provided they have not regrown from clearcutting, but may have been grazed or experienced limited exploitation of valuable tree species, and their floors may have been burned by Native Americans and European pioneers (Leverett 1996).

Historical criteria do not yet provide for indirect human disturbance, such as air pollution and introduced species, which are likely more pervasive than direct human disturbances (Runkle 1996).

Utilitarian Definitions of Old Growth

In some instances, so-called utilitarian attributes, such as stand age, are used to define old-growth. In addition to being easy to apply in the field using tree cores or historical records, stand age is also important on a human scale: one values old growth because it is hard to replace. Stand age also gives an objective measure of how long a particular forest will take to be replaced. However, using an arbitrary age as the defining factor of old-growth has shortcomings. Age is not based on the stand cycle, it may include stands of varying structure, and it is not clear how to treat minor disturbances (Runkle 1996).

Economic Definitions of Old Growth

Economic criteria have also been used to define old-growth. Using economic criteria, old-growth is reached when trees show signs of advancing age, such as crown die-back and heartwood decay, and tree growth rates are slowed. Forests falling into this category are considered to be "overmature," meaning they have reduced economic value. However, this method ignores other likely causes of tree decline that may occur in young forests, such as pollution and disease (Leverett 1996).

The Need to Consider Landscapes

Practical definitions of old growth tend to be based on tree characteristics and stand properties. But definitions of old growth can also be addressed at larger scales. For example, forested landscapes dominated by old growth can incorporate patterns of natural disturbance by including younger patches of recently disturbed forest interspersed in the older matrix. Different forest types that may occur close together due to soil or topographic gradients and slope aspect differences can be classified as an old-growth landscape, rather than separately defining adjacent stands of old growth that fall within different forest type groups (Tyrell and others 1998).

Researchers may answer the question of what to call old forests of relatively young trees by giving separate meanings to "primary" and "old growth." "Primary" would refer to forests of any age that have experienced no or only minimal human disturbance. On the other hand, "old growth" would refer to forests that meet certain criteria of age and/or stand development (Leverett 1996).

Tyrell and others (1998) proposed an operational definition for old-growth that provides a tool for identifying the degree of "old-growthness" that a particular stand exhibits. This old-growth concept may be diagrammed as a wheel with a central hub and radiating spokes which represent defining characteristics. There are 15 defining characteristics, hence 15 spokes. The scales are defined so that the closer a value is to the hub on each spoke, the closer the value of that characteristic is approximating the core concept of old-growth (Tyrell and others 1998).

Tyrell Model of Old-Growth Forests

The core concept for old-growth used for this model is:

"an old, uneven-aged forest with large trees and lots of dead wood, located on productive sites at low elevations, connected to other old-growth sites via forested corridors, and not likely to experience large-scale disturbance, embodies the quintessential old-growth concept" (Tyrell and others 1998).

The process of using this model to estimate the degree of "old-growthness" of any stand is straight forward. The target forest is assigned a value for each spoke (characteristic). This value is marked on the spoke of the wheel diagram at the appropriate place. All marks are then connected to produce a polygon. For example, compare an ideotypical northern hardwood forest with an ideotypical aspen forest. The smaller the polygon, the more closely the target forest resembles the core old-growth concept. The ideotypical aspen forest diverges from our core old-growth concept on several spokes (characteristics), thus resulting in a larger polygon.

Distribution and Description of the Mixed-Mesophytic Forest Type

Structure, Composition, and Dynamics of Old Growth

See: Standardized Table of Old-Growth Attributes for Western and Mixed Mesophytic Forests.

Diversity

Disturbance

Canopy Gap Formation

Age Structure and Diameter Distributions

Old-growth stands differ from younger, uneven-aged forests in having a greater range of tree sizes, a greater maximum tree age, and more large-diameter trees.  Few studies cite age exceeding 250 years for trees in old-growth, mixed mesophytic forests, although occasional trees much older than 250 years have been reported.  Irregular age distributions are common in old-growth stands and reflect severe natural disturbance or irregularities in seed production (Lorimer 1980). Estimates of background tree mortality for all species range from 5 to 10 percent per decade in old-growth, mixed mesophytic forests.  An additional 6 to 8 percent mortality caused by disturbance within a given decade is sufficient to create peaks in diameter distributions as more seedlings and saplings survive and grow into the canopy stratum (Lorimer 1980, Greenberg and others 1997).

Old-Growth Dynamics

Buckner (1989) described the pine-hardwood type as a "mid-seral stage that is ephemeral on a given site and is maintained in a changing landscape mosaic where scattered disturbances reinitiate succession in a stochastic manner."  Oak-pine forest types, as such, do not adhere to steady-state dynamics, and more traditional definitions of "old-growth."  White and Lloyd (1998), defined the old-growth oak-pine type as "resembling the transition or understory reinitiation phase of forest development or the mid-to-late phase of plant succession."  They include oak-pine stands that have minimal evidence of postsettlement human disturbance and contain pines exceeding 100 to 125 years old in their definition of old growth (White and Lloyd 1998).

Distribution and Composition of the Dry and Dry-Mesic Oak-Pine Forest Type

Dry and dry mesic oak-pine forests occur along dry ridges and south-facing slopes throughout the southern Appalachians.  By definition, dry and dry mesic oak-pine forests must contain at least 20 percent pine basal area and at least 20 percent oak basal area. The relative dominance of yellow pine species (Virginia, pitch, Table Mountain, and shortleaf) is determined by elevation, exposure, and disturbance regimes.  Scarlet oak and chestnut oak are the most common oak species of this forest type.  Other hardwoods typical of this forest type include: black and white oaks, blackgum, red maple, yellow-poplar, dogwood, sassafras, and sourwood.  Historically, chestnut was a component of the dry and dry-mesic oak-pine type (White and Lloyd 1998).

Although the abundance of the oak-pine type in the Appalachians at the time of European settlement is not clear, use of fire by Native Americans (Buckner 1989, Van Lear and Waldrop 1989), coupled with natural disturbances (wind and ice storms, insects, pathogens, and lightning fire) enhanced and maintained the oak-pine type in presettlement times in the Appalachians.  Native yellow pines (Table Mountain pine, shortleaf, pitch, and Virginia pines) all share some degree of fire tolerance. With increased fire suppression, yellow pines will become less abundant in these forests and, in some cases, oaks will eventually decrease as well (Nowacki and Abrams 1992).  If disturbance regimes are not restored, white-pine, a less fire-tolerant but more shade-tolerant species, may become an important component of old-growth oak-pine forests. Maples, black cherry, and black birch also will increase in the absence of fire. Plant diversity in most dry and dry-mesic oak-pine forests is relatively low, particularlty in the absence of fire when forest understories become dominated by one or two layers of ericaceous species (White and Lloyd 1998).

Structure and Dynamics of Old Growth

Information Gaps

• Locating, documenting, and describing existing stands and their old-growth characteristics.

• Documenting threatened, endangered, and sensitive plant and animal species that are associated (currently or historically) with this forest type.

• Researching presettlement forest conditions, spatial and temporal scales of nonhuman disturbances, and the distribution of Native American populations and their impact on presettlement landscapes.

Historical fire regimes throughout North America were greatly influenced by aboriginal man. Homo sapiens sapiens migrated to North America across the Bering Strait about 20,000 to 35,000 years ago (Aschmann 1978, Komarek 1974).  Ancestors of these original peoples arrived in the southern Appalachians about 10,000 years ago (Keel 1976).  Natural fire regimes were dramatically altered by paleo-Indians.  In the southern Appalachians, as elsewhere, paleo-Indians increasedfire frequencies from the natural fire regimes of mostly lightning ignited fires (Van Lear and Waldrop 1988).

When paleo-Indians first arrived in the southern Appalachian region, the landscape was dominated by boreal forests (tundra or taiga). Gradual global warming shifted the dominant forest type in the southern Appalachians to upland hardwood forest (Delcourt and Delcourt 1991).  It is believed that paleo-Indians initiated widespread burning to encourage grazing habitat in these deciduous forests.  Hunting was their primary means of survival for most of the millennia that they occupied the southern Appalachians (Buckner and Turrill 1999).

It is estimated that paleo-Indians in the Southeast developed agricultural techniques around 800 to 1000 A.D. (Hudson 1982).  Fire was used by these early agriculturalists to clear fertile floodplains for cultivation (Delcourt and Delcourt 1997, Chapman 1985).  Fires set intentionally by paleo-Indians in agricultural plots likely escaped to surrounding uplands.  But paleo-Indians also intentionally used fire outside of cultivated lands for other benefits including improving grazing habitat for wildlife, exposing nuts, and encouraging fruit production (Williams 1989, Buckner and Turrill 1999).

The relatively high human population densities in prehistoric America, although still a subject of much debate, indicate that most regions were likely subject to frequent anthropogenic fires.  Dobyns (1983) estimated that 18 to 20 million native Americans inhabited North American in 1492 (Dobyns 1983). Evidence from paleo-ecological studies also indicate that during most of the last 4000 years, paleo-Indians played an important role in determining the composition of southern Appalachian vegetation through their selective use of fire (Delcourt and Delcourt 1997).  Perhaps the best, and most objective, evidence about the composition of forests before European settlement comes from pollen records from pond and bog sediments that have accumulated for thousands of years.  These studies indicate that anthropogenic fires increased populations of fire-tolerant oaks, chestnut, and pines in upland forests of the southern Appalachians (Delcourt and others 1986, Delcourt and others 1998, Delcourt and Delcourt 1997).  Landscapes likely contained open pine and oak forests with widely spaced trees and herbaceous understories when European settlers arrived.  For example, the Shenandoah Valley of Virginia was reported to be a vast prairie between the Blue Ridge Mountains and the Alleghenies in the mid-1700s (Leyburn 1962, Van Lear and Waldrop 1988, Buckner and Turrill 1999).

Old-growth forests can provide many services and support many desirable uses.  Giles (2000) listed some important and obvious benefits:

• Aesthetic appreciation
• Education
• Recreation
• Preservation of genetic material for unknown future services
  • Potential for discovering medicinal plants
  • Potential for bioengineering
  • Source of potential inoculum for recovering forests
• Research of natural processes
• Watershed protection and runoff control
  • Protection from erosion
  • Reduces peak flows of floods
  • Filters and slows water
• Retention of plant and animal diversity
• Carbon storage and therefore a potential buffer of global warming
• Provides habitat for a wide variety of animal and invertebrate species
  • Supports a rich insect faunal community.
  • Provides habitat for rare and endangered animal species.
  • Provides a multi-layered, three dimensional habitat with complex structure.
  • Provides vertical structure to increase bird diversity.
  • Provides micro-"travel lanes," both fallen and standing trees for many animals.
  • Provides critical habitat for several species of reptiles and amphibians.
  • Provides critical habitat for several species of tree bats.
  • Produces abundant and varied mast for animal consumption.
  • Provides sources of mycorrhizal fungi as wildlife food and for invaluable tree-growth enhancement
  • Produces relatively large amounts of dead wood as habitat for arthropods that are a food base for birds and some mammals.
  • Provides perching and cavity sites, particularly for the larger birds and mammals (e.g., pileated woodpecker, bears, raccoons).
• Produces climatic modification due to changes in albedo and wind effects.

Many of these listed benefits are applicable to all forests, but old-growth may produce them in greater abundance.

Old-growth forests provide important habitat for a variety of species in the southern Appalachians. In particular, many of the taxa that contribute to the Appalachians high levels of biodiversity appear to be dependent on old growth. Mesic old-growth forests of this region contain extraordinary densities and diversities of both salamanders and vernal herbs. In contrast, secondary forests of the same type have severely reduced populations and diversities of both (Meier and others 1996).  For example, one study of Appalachian cove hardwoods definitively showed that species richness and relative abundance were lower in young-aged stands than in those greater than 85 years of age (Ford and others 2002).

Salamanders and vernal herbs share several characteristics that make them old-growth specialists. They depend on moist environments, have long generation times, low reproductive rates, and slow dispersal (Meier and others 1996). In addition, both salamanders and herbs benefit from the formation of canopy gaps, pits, mounds, and rotting logs-- features common in old-growth forests. For example, Bratton (1976) found that some species of vernal herbs most commonly root in deep pockets of organic matter at the bases of trees or on fallen logs. Also, salamander density and diversity have been shown to be correlated with the availability of well-decayed coarse woody debris (Meier and others 1996).  It has been shown that clearcutting Appalachian cove hardwood stands negatively impacts several salamander species and that recovery time exceeds 50 years (Ford and others 2002).

Given the close relationship of these taxa and old-growth, conservation of species diversity in the southern Appalachians depends in part on the preservation of existing old-growth forests.  Meier and others (1996) discuss several points for the conservation of species diversity in old-growth stands. First, research has indicated that even small remnant primary old-growth forest stands are important reserves of vernal herb diversity. Nevertheless, small tracts of old-growth forest may not be adequate to preserve diversity on regional and larger scales.  Both large and small blocks of old growth, therefore, should be protected where possible. Second, the low to nonexistent recovery rates observed for vernal forest herbs suggest that forest blocks restored to old-growth conditions cannot be depended on to conserve and restore vernal herb populations. Third, due to the poor dispersal characteristics of many vernal herbs and salamanders, fragmented old-growth blocks will be less effective in species conservation than a network of corridors between isolated stands (Meier and others 1996).

However, Ford and others (2002) argue it is not old-growth forest per-se that is important to salamanders but rather the existance of cove hardwood forest habitats.  In addition, it is the landscape connectivity of these cove hardwood forests that aids in species diversity, dispersion, and genetic integrity.  It may be somewhat dangerous to focus on any one particular forest condition, such as old-growth, as a surrogate for understanding the habitat needs of the species or species groups of interest.  These needs may be quite adequately met by forest conditions other than old growth.

The importance of old-growth forests to carnivores appears to be an open scientific question.  One theory, represented by Hunter (1990) and Pelton (1996), holds that old-growth forests and, by extension, low levels of active forest management, are positively correlated to viable and healthy carnivore populations.  In contrast, a second theory has been developing that maintains that "A well-managed forest that provides adequate opportunities for movements, foraging, denning, and cover has a better potential to satisfy the needs" of many carnivore species, especially black bear, than an unmanaged, old-growth forest (Weaver 2000).  Since theories have scientific credibility, we will summize the main points of both in this section.

Theory:  Old-Growth Forests are Very Important to Carnivores

Old-growth forests provide several important habitat features for carnivores, such as escape cover, maternity sites, and winter dens. Most carnivores are sensitive environmental indicators; their low population densities and high trophic level make them vulnerable to habitat perturbations. Carnivores, therefore, can serve as important umbrella or flagship species (Hunter 1990, Pelton 1996).

Carnivores benefit in a variety of ways from microhabitats in old-growth forests. Benefits include security, energy savings, permanence, psychological health, comfort, food, water, thermal stability, and flexibility in denning. For some species, the availability of a specific component of older forests may be crucial for survival. For other species or under other conditions, a specific component may only be facilitative. A brief review of the benefits of old growth for carnivores outlined by Pelton (1996) follows:

Benefits of Old Growth to Carnivores

• Security
  The reclusive nature of most carnivores necessitates seclusion at various points in their life cycle. Cavities-- particularly in large standing trees, dead or alive-- offer a high degree of protection from weather and other predators.
• Energy Savings
  Thermal exchange between a mammal and its environment is an important biological function affecting behavior, habitat selection, energy budget, reproduction, and survival. Bears denning in large trees were estimated to save 40 percent in energy compared with bears using ground dens.
• Permanence
  Old-growth sites exhibit a degree of longevity that far exceeds anything in disturbed or intensively managed forests. Dens in stumps, root masses, tree trunks, or logs may persist for more than 100 years. Persistence of such structural components in a landscape provides a degree of stability that could be beneficial to many carnivores that are often long-lived.
• Psychological Health
  Many carnivores are born in and remain in dark, vertically walled enclosures for extended periods. Such structures, therefore, may be an important part of their survival strategy throughout life and fulfill their psychological needs.
• Comfort
  The decay processes of older forests provide spongy, soft, and dry wood fragments that serve as excellent bedding material and ideal nesting material for newborn young.
• Food
  Although for many carnivores old-growth sites may not provide the quantity or even quality of food found on earlier successional sites, the old growth does represent a source of food stability and diversity as compared to intensively managed forests.
• Water
  Old growth can provide small pockets of water in the crook of large limbs, stump/snag depressions, and hollow stumps over extended periods when no other local sources of water are available.
• Thermal Stability
  Old-growth stands are generally moister and cooler than younger forests, providing optimal rest areas during the heat of summer.
• Flexibility in Denning
  Older forests offer a wide array of denning opportunities, including soil pits created by large root masses of wind-tilted trees, the root masses themselves, stumps, logs, large horizontal limbs, cavities in standing trees, and light-gap thickets created where large trees have fallen.
• Necessity
  Older forests may provide denning opportunities where no others exist. Older forests may become more important to carnivore populations when harvest intensity is high, either for the species itself or for its habitat. For some species, such as raccoon or black bear, the need for large tree refuges may be critical if the method of harvest of the population is pursuit with dogs rather than still-hunting.
• Preference
  Studies of black bears, raccoons, and striped skunks have demonstrated preference for various components of old growth. For example, in areas where potential ground dens are abundant, pregnant female black bears and raccoons select more protected high tree cavities (Johnson and Pelton 1981).

How Old Must Trees Be?

For raccoons, den trees averaged 60 cm d.b.h. and ages ranged from 90-164 years (Cantrell 1989). For bears, den trees averaged 100 cm d.b.h. and ages ranged from 175 to 280 years (Wathen and others 1983). Gray squirrels, prey for some carnivores, require 40 cm d.b.h. and larger trees ranging from 65 to 130 years (Nixon and Hansen 1987). Therefore, bears and raccoons prefer trees that exceed rotation ages for eastern forest stands (Pelton 1996).

Managing Old Growth for Carnivores

Old-growth stands should not be solitary. They should occur in large, well-distibuted blocks so that carnivores have a wide array of choices. Exactly how much old growth is needed should be addressed through intensive research-- particularly for species whose requirements encompass the needs of many other species (Pelton 1996).

Theory:  Old-Growth Forests are NOT Particularly Important to Carnivores

While old-growth forests may enhance important habitat features for one or more carnivore species, most carnivores are not obligatory inhabitants of old-growth forest or wilderness areas.  A well studied example is the black bear (Weaver 2000).  Far from supporting the crucial role of old-growth forests in carnivore population health, this theory claims that a well-managed forest has better potential to satisfy the needs of carnivore populations than an unmanaged, old-growth forest or wilderness.  It is much more important to understand species requirements and adaptabilities and then to design a forest management program that changes and maintains the landscape in a way that supports the health and vitality of that species as opposed to choosing one type of forest condition, such as old growth, as a surrogate for sound management.

For example, black bear habitat and landscape management in the central hardwood region of the United States includes the active use of prescribed fire as well as wildfire suppression (Weaver 2000):  Fire Management is needed to:

• Create and maintain habitat diversity.
• Provide stable, diverse, and widely available food supplies throughout the year.
• Maintain and create suitable denning sites and other habitat.
• Provide escape cover.
• Provide reforestation opportunities to maintain juvenile forest types.
• Manage human access and activities to decrease negative human-bear incidents.

In the mid 1980s, when forest plans were being written for the eastern U.S. National Forests, not one considered managing for old growth. Now some National Forest plans are calling for from 5 up to 30 percent of the land to be managed for old growth (Tyrrell 1996).  This approach may be largely attributed to the Endangered Species Act, which encourages forest managers to set aside areas of forest for the protection of certain endangered species.  However, these allocations are not for old-growth preservation.  Rather, they are extensions of rotation ages with the intent to harvest as stands begin to exhibit old-growth qualities (Zahner 1996).

An amendment to the management plan for the Nantahala-Pisgah National Forest in North Carolina calls for setting aside 32 separate "patches," each of 2,500 contiguous acres or more, of forest with old-growth management potential (USDA Forest Service 1994).  The affected area covers some 80,000 acres that may become older forest. But Zahner warns that only 8 percent of total forest land in these national forests will be affected and notes that these areas are not permanent old-growth reserves (Zahner 1996).

In the Southern Appalachian Assessment (SAMAB 1996), an inventory of national forest land was conducted which identified areas that might meet goals for old-growth management. Stands were selected based on three criteria: (1) forest stands older than 100 years, (2) forest stands included in designated wilderness areas, and (3) stands known to have received little disturbance. It should be stressed that the stands identified in the assessment may or may not be managed as old growth; rather, areas to be managed as old growth will be be selected from this group in future forest plan revisions (SAMAB 1996).

Distribution of Possible Old Growth
in National Forest System Lands in
the Southern Appalachians

The initial inventory identified almost 1.1 million acres, ranging in size from 1-13,000 acres, as potential areas for the management of old-growth forest. Eleven forest types were represented in these potential old-growth stands, with nearly half of the area in dry to mesic oak types. Other forest types with high representation were: mixed mesophytic; dry and dry-mesic oak-pine; dry-xeric and xeric oak; and, xeric pine and pine-oak. Almost 39 percent of these forests are currently being managed for timber production. Each stand can be viewed at SAAs online reports website.

Five of the Appalachian National Forests released revised management plans in January 2004 that implement a new old-growth management strategy.  The new strategy includes the designation of a network of large, medium, and small potential old-growth blocks, protection of existing old growth, and restoration of old growth through both active and passive means (USDA Forest Service 1997).

Current Condition of Eastern Forests

Areas of Known Old Growth in the Southern Appalachians

Great Smoky Mountains National Park (GSMNP), which straddles the border between North Carolina and Tennessee, is the largest reserve of old-growth in the southern Appalachians.  Of the parks 175,000 acres, an estimated one third is old-growth, distributed among the following forest types: cove hardwoods (38.7 percent); northern hardwoods (15.3 percent); mixed mesic hardwoods (13.3 percent); and other types ranked at below 10 percent each.  Virginias George Washington National Forest has approximately 42,000 acres of primary forest, containing mostly dry-mesic oak, and xeric pine and pine-oak. The extent of old growth in western North Carolinas Pisgah and Nantahala National Forests is still undetermined, but identified sites total at least 18,000 acres.  Old-growth in these forests are found in the Linville Gorge Wilderness Area, Big Ivy, and Joyce Kilmer Memorial Forest.  Joyce-Kilmer Memorial forest contains a cove-hemlock forest with trees believed to be close to 500 years old.  Although most of Tennessees known old-growth is in GSMNP, patches of old-growth are located in the Cherokee National Forest (>2,000 acres).  The only known old-growth in Kentucky is contained in the Blanton Forest (2,350 acres).  On the border of South Carolina and Georgia, the Ellicott Rock Wilderness area in the Sumter National Forest contains 1,000 acres of old-growth that is more than 90 percent upland oak.  Georgias Chattahoochee National Forest contains roughly 5,000 old-growth acres, found mostly in the Cohutta Ranger District (upland oak and pine) and the Chattooga watershed (oak) (Davis 1996).

Davis (1996) warns that her figures include primary forest in these figures that would not satisfy the definitions of old growth followed by most researchers.  She also warns that her figures are more a reflection of what we know, not what actually exists.  Note also that the acreages she included not all under protection.

Era of Fire Suppression

Federal land acquisition in the early 1900s initiated a change in burning practices throughout the southern Appalachians, and began an era of fire suppression. Eliminating all fire from public forests was the official stance of federal agencies in the early 1900s when the Forest Service was established (Dyne 1982). Foresters of that time did not realize the important ecological role of fire in the development and maintenance of the ecosystems they were trying to protect (Van Lear and Waldrop 1989). As both the USDA Forest Service and state agencies improved their capabilities for fire detection and suppression, fire was essentially removed as a vector for shaping landscapes in the southern Appalachians (Buckner and Turrill 1999). For example, the creation of the Great Smoky Mountains National Park (GSMNP) in the 1930s altered fire frequencies from once every 10-40 years (from the period of 1856-1940) to once every 2000+ years ( Harmon 1982, Buckner and Turrill 1999).

In the 1940s, a federal fire prevention campaign was initiated, partly in response to several catastrophic forest fires in the Great Lakes Region caused by poor logging practices. Smokey Bear became the highly effective symbol of this campaign, teaching two to three generations of Americans that all fires are harmful to forests (Buckner and Turrill 1999).

See: The Effects of Fire Suppression

Trombulak (1996) proposed some general principles for planning a landscape-scale approach for old-growth restoration and maintenance:

• Small preserves are more vulnerable to destruction than large ones.  Natural disturbances from wind, fire, and disease are more likely to destroy 10-20 ha stands than 100 + ha stands.
• Old-growth restoration strategies should de-emphasize goals which aim for a specific set of species as an end point because of the natural unpredictability in patterns of species colonization and establishment as well as the impact of anticipated climate change.
• Old-growth restoration plans should apply to the long term-- to centuries.  Long-lived individual trees take centuries to develop old growth characteristics.  In addition, the abiotic and biotic environment in which these trees live also requires these centuries to fully develop into an old-growth community.
• Old-growth restoration efforts should focus on the old growth community complex rather than on a few of its elements in isolation. Old growth restoration may require active management until the conditions for the successful functioning of natural regeneration have been established.  Common impacts that might prevent old-growth from successfully regenerating include: eroded soils, unnaturally high populations of herbivores (deer or domestic animals), exotic plant competitors, exotic insect pests and pathogens, fire suppression and timber harvesting practices that increase the threat to old growth from any of these factors.
• A landscape restoration plan must include an design for human use of the old-growth landscape.  Permissible, as well as prohibited uses, must be clearly spelled out.
• A professionally responsible old-growth restoration plan will avoid questionable practices, which, upon careful examination, fail to bring about the true goal.  Some questionable practices are (Trombulak 1996):
  • Selecting stands for old-growth restoration because they are economically unsuitable for timber harvesting.
  • Selecting primarily recently clearcut or other heavily modified stands for the restoration program when other choices are available.  These types of stands will take so long to develop old-growth characteristics that the plan effectively excludes old-growth from a landscape for centuries to come.
  • Selecting small, widely scattered stands for old-growth restoration.  These isolated patches often have too large a probability distructive disturbance events to ever make it to old-growth status.

Major challenges related to planning at the landscape scale are: (1) deciding which areas we are going to protect and, (2) deciding which of the selected areas are going to be managed by passive restoration and which by active restoration.

Passive management for old growth restoration is simple. In many cases, natural ecological processes and time are all that are required to bring old growth into being (Trombulak 1996).  Protection from extractive utilization by humans will likely allow most second-growth forests in the southern Appalachians to mature into old growth over a 150-400 year timespan (Trombulak 1996). Passive restoration of old growth on public land through the establishment of parks, wilderness areas, and research natural areas has been going on for a long time. Old growth restoration using passive management is likely to succeed if society is willing to accept whatever old-growth forest ecosystems result.

Unfortunately, the impact of past extractive utilization of the forest, climatic change, exotic disturbance agents (such as bittersweet and high density deer populations), and protection from the occurrence of historic fire patterns may make it unlikely that historically representative old-growth forests can be restored using passive management alone.  In such cases, various active management practices might be attempted.

Old growth definitions based on structural characteristics of forests imply that silvicultural techniques can be used to actively create or maintain these old-growth conditions (Tyrell and others 1998):

"Silvicultural manipulations can indeed create or accelerate the formation of some of the structural attributes of old-growth (Nyland 1997): produce large trees through crown release (see also Goodburn and Lorimer 1997, Singer and Lorimer 1997), maintain an irregular upper canopy, increase vertical structural diversity, create snags, mimic gap dynamics (see also Runkle 1991), and increase age diversity. Other practices, such as adding nest boxes or dead logs, provide habitat associated with old-growth (Runkle 1991). However, some old-growth characteristics prevail only with sufficient time. For example, it takes time for abundant coarse woody debris to develop even with silvicultural manipulations (Goodburn and Lorimer 1997), and for logs to reach advanced stages of decay (Tyrrell and Crow 1994)."

Many silvicultural practices might be used in an active old-growth restoration project (Trombulak 1996, Giles 2000):

• Removal of exotic species.
• Thinning selectively to restore natural species composition and increase individual tree sizes more rapidly.
• Prescribed burning to mimic wildfires.
• Selectively harvesting trees to create clearings and open gaps in the canopy.
• Extending rotation ages.
• Re-introducing wildlife flora and fauna known to be historically present but currently absent.
• Creating standing snags.
• Re-introducing of organic log structures to streams.
• Excluding horses.
• Managing of populations of herbivores, such as deer.
• Identifying and eliminate vectors for the introduction of exotic plants.
• Managing local fish populations.

The active application of silvicultural methods to restore old-growth forest conditions is controversial and still largely experimental (Trombulak 1996). The controversy centers on the contention that silvicultural management techniques do not produce the full array of old-growth characteristics and do not, therefore, result in old-growth (Leverett 1996). Therefore, until we improve our knowledge and predictive capabilities, Trombulak (1996) advises that active management for old-growth restoration should be limited to small fractions of an entire restoration project. The results should be monitored and analyzed and management should be altered on the basis of analysis (Trombulak 1996).

Proponents of old growth argue that existing relics of old growth, preserved as local public parks and recreation areas, are not "enough." Davis (1996) estimates that old-growth occupies less than 0.5 percent of the forest land in the Southeast. Zahner predicts this amount is not sufficient, particularly because most existing old growth occurs as small isolated fragments with little connectivity and virtually no long-distance migration corridors. Further, they argue that if old-growth protection was expanded to include all National Forests, the amount of old-growth would still be insufficient (Zahner 1996). However, in answering the question: "How much old-growth is enough?" we must first ask "Enough for what?"

Enough for What?

Trombulak (1996) presents several additional arguments for the protection and restoration of old growth:

1. The primary goal for conservation is the protection and restoration of biological integrity (Noss 1992, Noss and Cooperrider 1994). Integrity is a function of four biological conditions: (1) representation of all ecosystem types and their successional stages, (2) representation of all native species in natural patterns of abundance and distribution, (3) maintenance of ecosystem processes, and (4) responsiveness to change. As a distinct successional stage of forest ecosystems, old-growth needs to be restored in order to promote biological integrity (Trombulak 1996). Zahner (1996) also argues that the most significant value of old-growth forests is their contribution to biodiversity.
2. Many species across a wide range of taxonomic groups are dependent on or positively associated with old-growth forests. As the amount of old growth declines, habitat for these species also declines, putting them at both demographic and genetic risk of extinction.
3. Old-growth forests serve as controls for understanding the impacts of forest management practices. As the amount of old-growth declines, our ability to assess scientifically the appropriateness of ecosystem management, new forestry, or any other management philosophy declines.
4. Forests in a primeval condition are aesthetically more pleasing as a setting for a "nature experience" than are modified forests.
5. Humans have a responsibility to practice compensatory justice with regard to the old-growth destruction we have caused elsewhere.
6. The current generation of humans has an obligation to future generations to pass on a world undiminished from that which was passed to it.

How Much Old Growth is Needed? The Biosphere Reserve Concept

The biosphere reserve concept is based on theories for preserving biodiversity. Briefly, the biosphere reserve model (Hough 1988) argues for several rules of size and distribution, including:

1. Large reserves are better than small reserves.
2. A single large reserve is better than a group of small ones of equivalent total area.
3. Reserves close together are better than reserves far apart.
4. Reserves connected by corridors are better than unconnected reserves.

If a reserve is to be effective in preserving all levels of biodiversity, its core must be surrounded by a buffer zone of habitats similar to those in the core but not necessarily protected from all human development activities. Populations of plants and animals that migrate only through mature forest cannot be sustained in isolation.  Examples of such plants and animals are old-growth herbaceous plants, especially perennials with limited pollination and seed dispersal, most arachnids and nonflying arthropods, most gastropods, and many amphibians and reptiles. Genetic interchange through linkages from habitat to habitat must be provided by establishing ties between adjacent communities and providing migration corridors between isolated fragments (Zahner 1996).

Though conceptualized on different scales, The Wildlands Project and Lifelands Project both borrow ideas from the biosphere reserve concept and propose to expand and link existing and proposed reserves (Wild Earth 1992). At the bioregional scale, Noss and Cooperrider (1994) estimated that between 25 and 75 percent of the total land area will be required for preserves, buffer zones, and connecting corridors.  Odum (1970) reached a similar conclusion, recommending that 40 percent of the state of Georgia remain or be restored as natural areas (Zahner 1996).

How Much Old Growth Can We Realistically Protect for the Future?

Zahner argues that, in the eastern United States, nonpublic lands have little potential to preserve or restore adequate amounts of old growth.  Restoration and preservation of old growth, therefore, will have to be accomplished on Federal land.  National Parks, National Wildlife Refuges, and National Forest Wilderness Areas make up less than 1 percent of the eastern forested landscape. National Forests, contain the greatest potential for maintaining and restoring old growth.

Although forest processes such as gap dynamics can be maintained in reserves of 500 to 1,000 acres, this size is likely to be vulnerable to destruction by catastrophic natural events such as fires, tornados and hurricanes (Noss and Cooperrider 1994).  Zahner argues, therefore, that National Forests create core areas of several thousand acres, each surrounded by buffer zones of light utilization.  Zahner proposes a very minimum of 50 percent of managed forest be set aside for old-growth preservation or restoration. However, assuming this is possible, only 5 percent of forest land in the East would be allocated as old-growth, a figure which already includes land preserved in National and State Parks.

Five percent is far below Odums and Nosss estimates of the required amount of land needed to sustain biodiversity in the bioregions of the East.  Zahner, therefore concludes that, National Forests combined with National Parks, cannot fully protect biodiversity.

See also: How Much Wilderness is Enough?

Wilderness, as defined by Section 2 (c) of the 1964 Wilderness Act, is a congressionally designated area of undeveloped federal land which appears to have been affected primarily by the forces of nature, has not had permanent improvements or human habitation, possesses outstanding opportunities for solitude or a primitive and unconfined type of recreation, and is protected and managed to allow natural ecological processes to operate freely (FSH 1909.12, Chpt. 7, item 7.1) (SAMAB 1996).

Wilderness areas are not the only large tracts of natural appearing forest land in the southern Appalachians.  Roadless areas, defined as natural areas with no more than 0.5 mile of improved road for each 1,000 acres, identify forests where signs of prior human activity are disappearing and the area is regaining a natural appearance.  The management activities allowed on wilderness and roadless areas differ in both their type and intensity (SAMAB 1996).

Wilderness and roadless areas provide humans with both direct and indirect benefits. Most of these benefits derive from the fact that roadless and wilderness areas provide the majority of undisturbed forest patches within an otherwise disturbed landscape.  For species, processes, functions, and human values that depend on undisturbed forests, wilderness and roadless areas are a vital component of the southern Appalachian region. These many uses, services, and benefits of wilderness have either an economic or intrinsic value, albeit some of these values are more difficult to measure than others.

In its survey of the extent and location of wilderness and roadless areas, the southern Appalachian Assessment reported that wilderness and roadless acres account for 4 percent of all land in the southern Appalachians.  How much wilderness is enough?  There is little agreement on this issue.  Some claim we have already "locked up" enough land in the US, while others argue that additional wildland protection is necessary.

Management of wilderness

The primary management objective for wilderness, drawn from the 1964 Wilderness Act, is to allow natural disturbance processes to dictate change and manage change from human activities within defined limits.  For management purposes, the legally defined boundary of a designated wilderness is also considered the ecosystem boundary.  Some examples of management activities in wilderness areas include: campsite naturalization, wilderness education, trail maintenance and rehabilitation, removal of human-made structures, campfire bans, restrictions on type of use and limits on recreation use (SAMAB 1996).

Management of roadless areas

Roadless areas are not excluded from management activities; some roadless areas are managed to maintain their undeveloped character while others are available for resource utilization. Management actions in roadless areas could alter natural features so that all or portions of them no longer meet wilderness or roadless criteria.  When a proposed management action is located in a roadless area, Forest Service policy, in compliance with the National Environmental Policy Act (NEPA), requires that the effects of that action be evaluated to determine whether it significantly affects the roadless character of the area. This analysis is conducted with full public participation, and management actions may be declined by managers according to the results of the evaluation.  Some constraints on management activities in roadless areas include (SAMAB 1996):

• Timber harvesting, if more than 20 percent of the area is <10 years old.
• Road construction or reconstruction, if the resulting density of improved roads exceeds 0.5 mile per 1,000 acres.
• Planting non-native vegetation on more than 15 percent of the area.
• Constructing pipelines, transmission lines or utility corridors with cleared rights-of-way.
• Mining.
• Prospecting with mechanical earth-moving equipment.
• Building major recreation developments (campgrounds, visitor centers, etc).
• Changing recreation activities from nonmotorized to motorized.
• Constructing buildings or other structures.
• Building high-standard surfaced trails.

Recreation

Recreation is perhaps the most readily apparent value of wilderness. In 1993, the total number of visitor-days surpassed 300,000 for wilderness areas in the southern Appalachians (SAMAB 1996) (Table: Annual Recreation Use of National Forest Wilderness in the Southern Appalachians). Yet, the value of wilderness for recreation cannot be measured solely on the basis visitor days.  Recreation research has revealed that a high-quality outdoor recreation system requires a diversity of recreational opportunities; a concept operationalized as the Recreation Opportunity Spectrum (ROS) (Brown and others 1978, Brown and others 1979, Clark and Stankey 1979, Driver and Brown 1979).  Wilderness offers a unique type of recreation emphasizing naturalness, solitude, and freedom.  Psychological research indicates that outdoor recreation is goal directed; that is, people participate in outdoor recreation to satisfy certain motives.  Examples of motives closely associated with wilderness include enjoying nature, physical fitness, reduction of tensions, escaping noise/crowds, outdoor learning, independence, introspection, achievement, and risk taking.  Without wilderness recreation opportunities, people seeking to satisfy these motives may be unfulfilled (Manning 1989).

Wilderness also holds special value for more "pure" or highly developed forms of recreation-- recreation where the emphasis is placed on technique and setting without the distractions of technology or other societal intrusions.  Progression of recreational activity from novice to more specialized forms (such as whitewater rafting, mountain climbing, hunting, and fishing) often require natural, undisturbed environments which provide greater challenges for enhanced skills and experience.  These forms of recreation are sometimes called "wilderness-dependent" (Manning 1989).

See also: Use-Patterns of Wilderness and Roadless Areas

Science

Large natural areas, such as wilderness and roadless areas, are needed to provide laboratories for the study of natural ecosystem processes.  Wilderness areas provide excellent opportunities for scientific research for several reasons: (1) they contain whole drainages where land and water interactions can be studied on a range of scales. (2) They may contain animal populations whose entire range and habitat needs are met within the wilderness. (3) They are large enough to include a mosaic of vegetation types and ages on comparable sites. (4) They frequently provide excellent areas to study the natural background levels of environmental pollutants (Greene and Franklin 1988). In the words of one environmental writer, wilderness "holds answers to questions man has not yet learned how to ask" (Nash 1982). Wilderness areas also provide laboratories for social research that examines mans relationship with nature (Manning 1989). Evidence suggests that wilderness areas are indeed used extensively as natural laboratories.  Research and monitoring activities have occurred in at least 29 national forest wildernesses in the southern Appalachians (SAMAB 1996).

Culture

Many historians agree that wilderness has contributed to the distinctiveness of American culture.  One of the qualities that made America distinctive was the grandness and wildness of its nature.  Many of Americas first contributions to world culture celebrated its wilderness heritage.  Examples include poems by William Cullen Bryant, novels by James Fennimore Cooper, and landscape paintings by Thomas Cole, Frederick Church, Albert Bierstadt, and Thomas Moran (Manning 1989).

Some suggest that wilderness shaped not only Americas physical and mental image, but its personality as well.  Turner (1920) explains that pioneers experience in the wilderness of the American frontier marked them with a sense of independence, rugged individualism, and self-worth which defined a distinctive American personality.  Moreover, these characteristics were directly translated into our distinctive form of democratic government with its emphasis on maintaining personal freedom (Turner 1920, Manning 1989).

Historical Resources

Cultural resources such as archaeological sites, historic structures, and artifacts, have enormous importance for the preservation of traditional American values, of both native Americans and settlers.  Federal land managing agencies, regardless of their specific missions or mandates, are required by federal preservation laws (such as the National Historic Preservation Act, the National Environmental Policy Act, the Archaeological Resources Protection Act, and the Wilderness Act) to protect cultural resources on managed land (Neuman and Reinburg 1988).  Wilderness areas not only protect cultural resources, but preserve these resources in a natural setting (Flamm 1989).

Therapy

Wilderness has long been thought to have both physical and mental therapeutic value.  Early thinking on this matter was influenced by Sigmund Freud and the developing science of psychology, which suggested that mental dysfunctions were often caused by repressed desires forced upon us by the constraints of society (Nash 1982).  Wilderness, proponents argued, provided an opportunity to release those constraints and play out emotion and instincts (Marshall 1930, Manning 1989).

The therapeutic values of wilderness have received considerable attention in recent decades.  A substantial industry has evolved around these potential values, led by Outward Bound and the National Outdoor Leadership School.  Studies of the therapeutic values of wilderness use has revealed beneficial effects on participants self-concept or self-perception (Burton 1981).  Self-concept-related preference items, such as self-confidence, are consistently found to be important to a large number of wilderness visitors (Driver and others 1987).  Despite methodological shortcomings in many wilderness therapy studies, a growing body of evidence suggests that various therapeutic benefits from wilderness are real (Manning 1989).

Aesthetics

Aesthetics is another area in which wilderness has been subject to considerable reinterpretation.  For example, mountains were once considered as "ugly deformities on the Earths surface" (Nash 1982).  The Enlightenment, and later the Romantic movement of the 17th and 18th centuries, developed a more sympathetic and appreciative view of nature, in which wilderness was considered beautiful due to the awe, power, and sometimes terror, it signified within us (Nash 1982, Manning 1989).

Marshall (1930) later developed a more sophisticated philosophy of wilderness aesthetics, recognizing that: (1) nature is detached from all temporal relationships in that it is not rooted in any one period of human history; (2) it has an encompassing physical ambience in that we can be literally surrounded by its beauty;  (3) it has a dynamic beauty as it is always changing; (4) it has the potential to gratify all of the senses in that it can be seen, heard, smelled, tasted, and felt; and (5) it provides the best opportunity for pure or perfectly objective aesthetic enjoyment in that it is not created or affected by man (Manning 1989).

There is considerable evidence of the aesthetic value of wilderness today, including the millions of photographs taken by visitors to national parks, and millions of nature calendars and picture books published by environmental groups and others.  In the Recreation Experience Preference scales, the scale item "scenery" ranks as one of the most important motives of wilderness visitors (Driver and others 1987, Manning 1989).

Intellectual Freedom

Nash (1982) argues that wilderness is the ultimate source of intellectual freedom or creativity.  Using the writings of several natural philosophers, Nash suggests that wilderness provides the purest form of objectivity from which original thoughts might be derived, inspiring intellectual creativity and diversity.  Wilderness-inspired intellectual freedom has been found in several religious, artistic, and political movements.  The Puritans came to the wilderness of the New World to find spiritual freedom just as the Mormons went to the deserts of Utah.  Thomas Cole and his followers found artistic inspiration in wilderness, and more recently, Abbey writes that wilderness may someday be needed "not only as a refuge from excessive industrialism but also as a refuge from authoritarian government and political oppression" (Abbey 1968, Manning 1989).

Spiritual Values

Nature is such a powerful and universal element of our world that its relationship to things spiritual is inevitable. Symbolic of this relationship is the fact that the word "wilderness" appears nearly 300 times in The Bible (Nash 1982).  However, wilderness has been subject to conflicting spiritual interpretations. In conservative religious doctrine, wilderness was seen simply as a force to be controlled and conquered or as a storehouse of raw materials for mans exploitation (Manning 1989).

More recently, nature, and its ultimate expression as wilderness, has benefited from more favorable interpretations.  In the 19th century, Emerson and Thoreau formulated their philosophy of transcendentalism, suggesting nature as a setting or metaphor for higher spiritual truths.  The transcendentalist interpretation of nature was eagerly accepted by wilderness enthusiasts, notably among them John Muir, and today the transcendentalist tradition continues.  Environmental degradation is often described as "desecration," a term with obvious religious overtones (Manning 1989).

Spiritual values and personal introspection are often cited as important motives for people who visit wilderness areas (Driver and others 1987).  In fact, some have even suggested that wilderness preservation might be justified on the constitutional basis of maintaining religious freedom (Graber 1976, Manning 1989).

Large metropolitan areas are obvious sources of visitors to wilderness.  However, there is no definitive evidence that the proximity of wilderness to a large metropolitan area is a major factor influencing the amount of use it receives.  For example, the high-use Cohutta Wilderness is outside the 50-mile radius of Atlanta, GA, where studies that indicate the majority of use originates, but it is within the 50-mile radius of Chattanooga, TN, where little use originates.  Also, the Cohutta is contiguous to the Big Frog Wilderness, which receives low use that originates from the local area (Carlisle 1992).  Most likely, a combination of factors account for an wilderness areas level of use.  Some factos have been shown to influence use: 

• Long-established wilderness areas receive high use.
• Prohibitions on campfires and overnight stays limit visitor use.
• Scenic features, special destinations, or streams draw users.
• Publicity increases use.
• Proximity to smaller cities such as Asheville, NC, and Roanoke, VA increases use.
• Location of a wilderness within a larger, well-known unit of a National Forest or National Park increases use (SAMAB 1996).

Biological Diversity

Continued human manipulation of the environment has repeatedly failed to maintain ecosystem, species, and genetic diversity (Flamm 1988).  Wilderness areas, as some of the last reservoirs of natural land in the United States, are becoming increasingly important for the conservation of biological diversity (Davis 1988, Flamm 1989, Schonewald-Cox and Stohlgren 1989).  The large size of wilderness areas (usually >5,000 acres) allow protection of biodiversity at both large and small scales. Ecosystem diversity is maintained when natural ecosystem processes are allowed to act on large scales.  Species diversity is maintained if forests are protected at an habitat scale, permitting species to express behavioral and other phenotypic variation not otherwise possible (Schonewald-Cox and Stohlgren 1989).  Wildlands are particularly important for the protection of disturbance-sensitive species that may require vast expanses of undisturbed forest for all or part of their lives. Finally, wildlands play a critical role in genetic conservation. Gene banks, gardens, and zoos are a limited means of genetic conservation; in situ methods provide the only reasonable, cost-effective solution at a broad scale (Flamm 1989).

In the southern Appalachians specifically, 7 of a total of 10 ecosystem sections are represented in either wilderness or roadless areas (Table: Ecosystem Section Occurrence in Wilderness and Roadless Areas in the Southern Appalachians).  Nineteen federally listed threatened and endangered species (8 plant, 11 animal) occur or have occurred in 16 roadless areas.  They include peregrine falcons, carolina flying squirrels, shale-barren rockcress, red-cockcaded woodpeckers, Indiana bats, spruce-fir moss spiders, mountain golden-heather, swamp-pink, and spreading avens.  The roadless area in the Great Smoky Mountains National Park contains the largest concentration of federally listed species (9 species).  Approximately, 92 percent of wilderness acres and 95 percent of all roadless acres are classed as potential bear habitat (SAMAB 1996).

Watershed Protection

Both research and experience have demonstrated the values of undisturbed forests in providing high-quality water at a low cost. In fact, watershed protection was one of the central aims of the acquisition of eastern national forests under the 1911 Weeks Act.  The benefits to humans from untainted, naturally-flowing, and healthy watersheds cannot be overstated. Cleansing water, supporting complex wildlife systems, trapping and modifying chemical concentrations, recharging underground systems, regulating annual water regimes, recreation for man, and aesthetics are services provided by watersheds.  Wilderness continues to offer the surest long-term protection of watersheds at the lowest cost; it relieves humans from attempting to artificially duplicate literally hundreds of within-system functions (Flamm 1989).

Possible Old Growth

Approximately 48 percent of the 1,098,491 acres of possible old growth on national forest lands in southern Appalachians is located in roadless areas and wilderness.  All acres within wilderness are considered possible old-growth, an area representing 32 percent of total national forest old-growth in the southern Appalachians.  Old-growth is present in 125 of 139 roadless areas, but for most of these areas less than half their total area is old-growth (SAMAB 1996).

Geological Features

Geologic and land form features store evidence of the historical forces that shaped present-day ecosystems; their preservation not only helps to maintain landscape and ecosystem diversity but can also provide insight into past environments and clues to the evolution of living organisms.  Wilderness designation provides protection for both outstanding and representative natural geologic features (Flamm 1989).

Economic Value of Wilderness

The many uses, and benefits wilderness provides society have an economic value, but some of these services are more easily evaluated than others.  Wilderness recreationists incur costs for travel and equipment.  These costs are considered a minimum measure of recreational value because recreationists may be willing to pay more than required.  Sorg and Loomis (1984) reviewed a number of willingness-to-pay studies of wilderness recreationists and found that most values ranged from $13 to $20 per activity-day (1982 dollars). The high-quality water that flows off protected watersheds and is used for drinking and irrigation is another example of a tangible benefit of wilderness (Manning 1989).

Other wilderness values are less amenable to economic valuation.  For example, the vital ecological "services" such as clean air and climatic stability, the future usefulness of biotic and genetic diversity, and the therapeutic, cultural, intellectual, esthetic, and spiritual values of wilderness.  It has also been suggested that wilderness may have unique values involving the preservation of natural environments.  By preserving wilderness and avoiding the irreversible decision of development, we may be creating and capturing option, existence, and bequest values.  In other words, wilderness remains available as an option for those who do not now use wilderness but may wish to do so in the future (Krutilla 1967, Manning 1989).

Intrinsic Value of Wilderness

A relatively new school of thought suggests that nature has an intrinsic value, and therefore preservation of wilderness is an expression of mans moral and ethical obligation to the environment.  Aldo Leopold recognized man as part of a larger ecological community-- a community that should have moral and ethical rights just as our human communities (Leopold 1966).  Contemporary reinterpretations of Judeo-Christian teachings have also extended our moral and ethical responsibilities to the natural world, arguing that the scripture contained in Genesis may suggest Mans dominion over other life forms are more appropriately interpreted as expressing a stewardship responsibility rather than indiscriminate use (White 1967).  The "rights" of nature have even been tested in the court system.  In the landmark case of Mineral King Valley, California, Supreme Court Justice Douglas wrote that wilderness had a right to legal standing in the court (Stone 1974). His, however, was a minority opinion.  These changes suggest that the environmental movement is evolving from its "shallow" anthropocentric traditions to a "deep" biocentric philosophy (Manning 1989).

Wilderness and roadless acres account for 16 percent of federal land, 19 percent of national forest land, and 4 percent of all land in the southern Appalachians. Of the 10 ecosystem sections found in the southern Appalachian Assessment (SAA) area, 7 are represented in either wilderness or roadless areas (SAMAB 1996) ((Table:Ecosystem section occurance in wilderness and roadless areas in the southern Appalachians.)).

Wilderness

Wilderness areas account for about 1 percent of land in the Southern Appalachians. They are distributed across 39 areas containing a total area of 428,545 acres. All occur on national forest land except for one area in the Shenandoah National Park in Virginia. The Chattahoochee has the largest number of wilderness acres of any national forest (114,789 acres) (SAMAB 1996).

See: (Table:Acres of officially designated wilderness in the southern Appalachian Mountains)

Roadless areas

Wilderness and Roadless areas
in the Southern Appalachian

Roadless areas account for about 3 percent of land in the southern Appalachians. The 144 such areas comprising a total area of 1,231,961 acres. Although the majority of roadless areas occur in national forests, the largest roadless area identified (464,544 acres) is in the Great Smoky Mountains National Park (GSMNP). Only 4 roadless areas are in state parks. There are 139 roadless areas on national forests in the southern Appalachians. They include 752,654 acres and contain 61 percent of all roadless acres. Sixty-three national forest roadless areas are larger than 5,000 acres. The George Washington National Forest contains the largest amount of roadless acres at 261,985 acres or 35 percent of all roadless acres on national forest land in the region (SAMAB 1996).

Future Trends

The acreage of roadless areas in the southern Appalachians is expected to change as these areas are designated as wilderness areas or allocated to multiple-use management. It is difficult to predict future designations since such evaluation involves the public and Congress. It is likely that the area of wilderness will increase, even if only slightly, as roadless areas are designated as wilderness (SAMAB 1996).

Is the number and size of wilderness areas sufficient to protect the values and ensure the benefits they provide? There is little agreement on this issue: some claim we have already "locked up" enough lands in the United States, while others argue that additional wildland protection is necessary. Flamm (1989) presents several arguments for why we need more wilderness areas.

First, to protect diversity at both large and small scales, a full range of functioning, intact ecosystems should be protected. Nation-wide, many of these ecosystems are under-represented with respect to both their size and quantity. In the southern Appalachians, for example, 3 of 10 ecosystems are not found protected within wilderness (Table: Ecosystem Section Occurrence in Wilderness and Roadless Areas in the Southern Appalachians). Flamm also notes that simply saving "representative" ecosystems is not sufficient. Species are not randomly and uniformly distributed throughout an ecosystem; therefore if only one of several like ecosystems is preserved, entire species or key genetic variants could remain unprotected. Replicating similar ecosystems provides a safety net for little known species, some of which could be seriously endangered (Wilcove and Flamm 1986).

Second, size may be the most important factor in conserving biological diversity. Island biogeographic theory estimates that a tenfold decrease in area corresponds to a halving of the equilibrium number of species present (May 1981). Using this estimate, we can expect a significant loss of species dependent on wildlands in the southern Appalachians, considering that only 1 percent of the land area is protected as wilderness. Flamm suggests that greater attention be paid to habitat fragmentation, which effectively reduces the size of wilderness areas. Habitat islands in a fragmented landscape may be unable to support viable populations of certain species, particularly species with large area requirements- often the very species in need of protection (Norse and others 1986).

As with proponents of old-growth, Flamm (1989) argues for larger and more connected wilderness areas. His "macro-wilderness" areas, are similar in theory to The Wildlands Project and Lifelands Project, both based on the BROKEN-LINK biosphere reserve concept. However, the future projection for wilderness may be similar to that of old-growth. In a best-case-scenario, if all roadless areas were designated for protection as wilderness, only 4 percent of land area in the southern Appalachians would be represented.

See also: How Much Old Growth is Enough?

Although forests in the National Wilderness Preservation System are among the most well-protected land areas in the United States, these areas are not immune to human threats.  A survey of wilderness unit managers throughout the United States by Peine and others (1989) revealed a wide variety of  internal and external threats to wilderness.

Internal Threats: Recreation Related Impacts

• Damaged vegetation along trails, around campsites, or near waterbodies caused by recreationists or pack-animals.
• Compacted soil along trails, around campsites, or near waterbodies caused by recreationists or pack-animals.
• Litter, human or pack-animal waste, water pollution.
• Crowding on trails or campsites and conflicts between users.
• Trailhead vandalism, theft, boundary-related problems.

External Threats

• Exotic plants, pathogens, insect pests, wildlife species.
• Water, air, and noise pollution (from such things as motor vehicles, and military overflights).
• Livestock grazing.
• Logging.
• Mining and prospecting for minerals, thermal mining, oil and gas mining.
• Accidental forest fires.
• Poaching.
• Industrial development, urban encroachment.
• Overpopulation of native species.
• Incompatible use of inholding properties.
• Roads (including utility access) and railroads.
• Global climate change.
  

The importance of individual threats to wilderness vary greatly among different regions of the United States.  In the Great Smoky Mountains National Park (GSMNP), air pollution is the primary threat. Other external threats of key importance in GSMNP are tourism-related development adjacent to the park; poaching of plants and animals; fires set by man; and exotic pathogens, insects, and plants. The next wave of extreme stress will be from the exotic gypsy moth.  Peine and others (1989) noted that vegetation in the eastern United States tends to be more resilient to human-caused impacts than vegetation in the West.  For example, after a 50 percent decline in backcountry overnight use at GSMNP from 1980 to 1985, mapping of bare ground at backcountry campsites revealed a significant recovery of vegetation.

Peine and others (1989) also noted that perceived threats may vary among federal agencies. For example, the National Park Service more frequently cited threats associated with exotic species and poaching, while the USDA Forest Service was more concerned with crowding. The ability of managers to protect wilderness against these threats will also vary considerably among different types of threats. Contending with external threats is problematic since the means to deal with them tend to be beyond the authority of the unit manager.  Threats associated with recreational use, therefore, are most likely to receive aggressive management action.

Due to these pervasive threats, Peine and others (1989) warn that the concept of "permanent " wilderness might be optimistic. They argue that threats to the wilderness are numerous, not well documented, and managers are far from understanding the ramifications of these threats on the ecosystem processes which constitute the essence of wilderness values they are charged to protect.

Peine and others (1989) made the following recommendations addressing this these threats:

1. Establish long-term, ecologically based monitoring programs to evaluate natures response to anthropogenic influences.
2. Establish more uniformity of wilderness management policy among agencies.
3. Create a standard database on wilderness.
4. Improve communication between site managers and perpetrators of external threats.

Estimates of the Earths genetic reservoir vary between 5 and 80 million species.  Significant controversy surrounds these estimates.  For a global perspective on the importance of biological diversity explore the website of the Convention of Biological Diversity.

In the temperate zone, the southern Applachian region is  considered a center of biological diversity (Meier and others 1996, Boone and Aplet 1994).  It is difficult to estimate the total number of species that occur in the southern Appalachians because only a fraction have been documented.  The All Taxa Biodiversity Inventory in the Great Smoky Mountain National Park (GSMNP) hase identified more than 200 species new to science and projects that more than 100,000 species may occur within GSMNP alone.

Regarding biodiversity as a resource, in and of itself, is a concept that has emerged only since the early 1990s.  This new recognition has encouraged thought on how biodiversity is defined, how it is measured, how natural variables affect it, what factors threaten it, and its role in ecosystem functions.

Actively managing for biodiversity is also a relatively new idea. Yet, virtually all management actions, including the decision not to take action, can affect biodiversity at several different scales.  Therefore, forest managers need to be aware of both the positive and negative impacts of management activities on biodiversity.

See also: Importance of Old Growth to Biodiversity

Many people believe biodiversity only refers to the total number of species in a given area, but the idea of "biodiversity" encompasses much more than species richness. Putz and others (2000) define biodiversity as

"the natural variety and variability among living organisms, the ecological complexes in which they naturally occur, and the ways in which they interact with each other and with the physical environment" (Noss 1990, Putz and others 2000).

They distinguish this natural variability from those formed under the influence of human-mediated species introductions and substantially human-altered environmental processes and selection regimes (Bailey 1996, Noss and Cooperrider 1994).

Biological diversity can be expressed at various scales (including landscapes, ecosystems, communities, species, genes) and with different parameters representing structural, compositional, and functional attributes. The following table, adopted from Putz and others (2000), defines biodiversity at each of these scales for each parameter.

The following overview of methods used to assess biodiversity, summarized from Huston (1994), largely refers to one definition of biodiversity: the composition of species at the community scale. There are several other types of biodiversity, defined at different scales for both functional and structural characteristics. However, methods for assessing these components of biodiversity are not addressed here.

Components of Biodiversity

The concept of diversity has two primary components: richness and evenness. Richness is the number in a sample, sometimes referred to as density when the sample size is expressed in terms of area. Evenness is the relative number; the more equal the numbers, the greater the evenness. Richness and evenness can be evaluated at any scale, from the number of alleles for a particular genetic locus to the number of different landscape types on a satellite image. Diversity is quantified using statistical formulas that combine both richness and evenness (Huston 1994).

Biodiversity can also be categorized according to how it is expressed at different spatial scales. Alpha-diversity, a measure of within-habitat diversity (Whittaker 1960), reflects coexistence among organisms that are interacting with one another by competing for the same resources or otherwise using the same environment. Alpha-diversity is measured as the number of species (or other components of diversity) within an area of given size. Beta-diversity, a measure of between-habitat diversity (Whittaker 1960), reflects how organisms respond to environmental heterogeneity. Beta diversity is usually expressed in terms of (1) a similarity index between communities or (2) of a species turnover rate between different habitats in the same geographical area, often along some sort of gradient. High-beta diversity is the result of low similarity between the species composition of different habitats or different locations along a gradient. Gamma-diversity is a measure of large-scale biodiversity that reflects evolutionary rather than ecological processes. Whittaker (1960) defined gamma-diversity (or "geographical" diversity) as simply the number of species within a region, analgous to alpha-diversity but at a regional scale. Cody (1986) defines gamma diversity slightly differently as a regional-scale analog of beta diversity. Cody also uses gamma diversity to to compare species composition among similar habitats in different geographical areas, rather than between dissimilar habitats in the same geographical area. This approach can reveal that ecologically similar but taxonomically different species are performing the same role in similar communities separated by a given distance (Huston 1994).

Statistical Methods of Assessing of Biodiversity

There are many methods for collecting, analyzing, and presenting data related to biological diversity. Depending on the particular hypothesis, any one of many sampling designs and statistical analyses can be appropriate (Huston 1994).

Indices of Species Diversity

Shannon Index: H = -sum (p_i log p_i)

where p_i is the proportion of the total sample (the total number of individuals or total biomass) composed of species i. (Shannon and Weaver 1949)

Because Simpsons Index (Simpson, 1949) is actually an index of dominance, and thus tends to be inversely related to evenness and richness, it is often expressed as a diversity index, Simpsons D, of the form:

D = 1- lambda

Indices of "Pure" Evenness

These indices are ratios of a composite index to a theoretical index indicating complete evenness:

Species-Areas Curves

Species-area curves graphically illustrate how the number of species increases as the area sampled increases. Species-area curves can be used to determine the minimum sample size or area needed to adequately characterize the species diversity of a particular community. These curves can also be used to compare the rate of diversity increase with area, which may allow testing of carefully chosen hypotheses. Changes in the rate of diversity increase with area suggest a transition in the mechanisms that influence diversity from one set of factors at small spatial scales to a different set of factors at larger scales. The scale of environmental heterogeneity is one of the most likely explanations for patterns in species-area curves (Grieg-Smith 1957). Species-area curves can be described by the equation:

Affinity Analysis

Affinity analysis is examines spatial patterns of diversity on landscapes. It incorporates the concepts of alpha, beta, and gamma diversity (Scheiner and Istock 1987). Affinity analysis is a statistical analysis of a data matrix of sites (columns) and species (rows). This method compares the actual patterns of similarity among samples from a landscape with an expected similarity based on computer simulations that randomly sample the original data (i.e., bootstrapping) (Huston 1994).

Issues Concerning Diversity Indices

Huston (1994) brings up several concerns about the use of statistics to described biodiversity.

• Diversity statistics differ primarily in the degree to which they emphasize species richness versus species evenness. Hence, depending on the question being evaluated, one index may be more appropriate than another. For example, in situations where changes in relative abundance indicate the effect of an important process, it is appropriate to use an index, such as Simpsons D., that emphasizes evenness more than richness.
• Rare species are unlikely to be detected in small samples and therefore tend to be undersampled. The consequences are that the evenness of species abundances will almost always be overestimated and diversity and evenness will tend to be highly correlated. Thus, most diversity indices are likely to produce the same ranking of diversity, except for a few extreme situations.
• Several statistics should always be provided to represent a sample. At a minimum, the number of species in a sample of a specific size and a measure of the species evenness should be provided.
• Regardless of the statistics that are chosen to describe diversity, it is critical that the sample be collected using a statistical design that will allow a reliable estimate of community properties that are relevant to the diversity issue being studied.

Factors to Consider When Sampling Biodiversity

In general, the sampling design for the collection of diversity data is far more important than which of the many possible diversity statistics are ultimately used to describe and analyze the results. The most important considerations are the types of organisms being sampled and the spatial or temporal scale (Huston 1994).

In order to understand the mechanisms influencing biodiversity of a particular system, the organisms included in a diversity sample should be divided into functional classes. Diversity among different bird guilds (frugivores, granivores, carnivores, or insectivores) may be influenced by different processes and therefore should be analyzed separately. Likewise, plant life forms, such as woody and herbaceous, trees and shrubs, geophytes and epiphytes, etc., should be considered separately. The same rule applies to studies of diversity at larger scales, such as trophic diversity, guild diversity, and life form diversity (Huston 1994).

Organisms should be sampled at spatial and temporal scales that reflect the physical and biological factors that influence biodiversity. Such factors could include the spatial and temporal variability of resources, variation in climatic conditions, and the spatial distances and time spans over which organisms interact. The relationship between sampling scale and the processes that influence species diversity is the basis of the distinction between within-habitat diversity and between-habitat diversity (alpha and beta diversity) (Huston 1994).

No single process or theory can explain a phenomenon as complex as biological diversity.  Understanding diversity requires understanding the interactions among many ecological, evolutionary, geological, and biogeochemical processes.  Huston (1994) outlines a number of factors, listed below, that have been correlated with patterns of biological diversity.  Examples are an areas latitude, altitude, productivity, spatial heterogeneity, size, age, and disturbance regime. It should be stressed, however, that notable exceptions to the general patterns exist for each of these factors, and therefore general conclusions are difficult to determine (Huston 1994).

Latitudinal Gradients

For most groups of terrestrial plants and animals, diversity decreases with increasing latitude.  Diversity is highest in the tropics and lowest at the poles. There are many factors correlated with this gradient that may affect species diversity, such as average temperature and precipitation, the variability in temperature and precipitation, annual net primary productivity, and geological history. Among the mountain systems of the eastern United States, the southern Appalachians are considered to be the center of diversity for a number of taxa (Huston 1994).

Altitudinal Gradients

Species diversity generally decreases with increasing elevation.  Many physical conditions, such as air temperature, rainfall, wind patterns, and the variability of those conditions, change along altitudinal gradients. For example, air temperatures, in general, decrease 6^oC with a 1000 m increase in elevation (Holdridge 1967).  Altitudinal gradients may be a large influence on patterns of biodiversity in the southern Appalachians; the region ranges in elevation from roughly 500-6,400 feet.  The most notable pattern in biodiversity is the change in community composition with altitude.  For example, spruce-fir forest types occur above 5,500 feet in the southern Appalachians while mixed mesophytic hardwood forests occurs at low to moderate elevations (Huston 1994).

A gradient study currently underway at Coweeta Hydrologic Laboratory is examining changes in a variety of ecosystem factors along an elevational gradient.  Preliminary results suggest that local topography, soil, and vegetation mediate, to some extent, macroclimatic gradients caused by elevation.

Productivity

Primary productivity, the solar energy that is captured by plants and converted to carbon compounds, has been correlated with biodiversity in many instances.  Precipitation, length of growing season, potential evapotranspiration, air temperature, and solar radiation, have all been demonstrated to be strongly correlated with species richness for a variety of taxa (Huston 1994).  Abundant rainfall is often cited as one of the factors contributing to the southern Appalachians high level of biodiversity (Huston 1994).

Diversity and the Size of the Sample Area

Large areas generally have more species than small areas.  Huston (1994) provides three potential explanations for the general increase of diversity with area:

1. It is a sampling artifact. Within a homogeneous area, samples of increasingly larger area will randomly sample an increasing proportion of the total population, and are thus likely to detect increasingly rarer species as the size of the sample increases.
2. It results from an equilibrium between extinction and immigration, or the number of new species immigrating to an area is balanced by the number of species emigrating or going extinct (as described by the equilibrium theory of island biogeography in  MacArthur and Wilson (1963).
3. It is a result of environmental heterogeneity. Increasing the sample area includes additional habitat types with groups of different species. Diversity will increase until all habitat types are sampled, which could require sampling of the entire Earth if the study were not limited to a particular geographic area (Huston 1994).

Diversity and the Spatial Heterogeneity of the Sample Area

Spatial heterogeneity is, in general, positively correlated with biodiversity (Huston 1994). There are many causes of environmental heterogeneity, including those acting on large scales (geology, climate, topography, and major disturbance) and small scales (vertical complexity created by plants, and minor disturbance).  One of the classic generalizations of ecology is based on MacArthur and MacArthurs (1961) observation that bird species diversity is positively correlated with the structural complexity of the vegetation. In the southern Appalachians, topography creates high spatial heterogeneity, and community composition often differs between north- and south-facing slopes, coves and ridges, and riparian and upland areas (Huston 1994).

Diversity and the Age of the Sample Area

Age, the time over which living organisms have been continuously present in an area, is correlated with biodiversity over temporal scales ranging from days to millions of years.  Three primary processes influence the diversity/age relationship:

1. Dispersal and migration: the number of organisms potentially present in an area increases with time as more and more species arrive;
2. Biologically produced changes in heterogeneity; and
3. The balance between speciation rates and extinction rates.

The high diversity found in the southern Appalachians is often partly attributed to its age; geologically, the southern Appalachians are one of the oldest mountain systems in the world (Huston 1994).

Diversity and Disturbances

Disturbance is correlated with biodiversity on many different scales, with some positive correlations and some negative correlations (Huston 1994).  Infrequent, massive disturbances such as glaciation, can lower diversity by killing or driving off organisms and altering the landscape. On the other hand, frequent and less severe disturbances have often been shown to increase biodiversity.  In fact, many studies have found that the highest species diversity occurs at intermediate frequencies of disturbance, with low diversity at both very high and very low frequencies, a pattern conceptualized in the "Intermediate Disturbance Hypothesis" (Connell 1978, Huston 1994).

Diversity and Biological Interactions

Biological interactions, such as competition, predation, mutualism, parasitism, and disease, also influence species diversity, but the importance of these interactions varies greatly.  In stable and productive environments, the effect of biological interactions on species diversity tends to be stronger than in unstable or unproductive environments (Szaro 1996).

Temporal Variation in Species Diversity

Diversity is constantly changing at many temporal scales.  Fluctuations in species diversity are a natural phenomenon in all ecosystems.  They occur in response to seasonal cycles, long-term climatic variations, and ecological processes such as succession.  In some groups of organisms, diversity measured at one time of year is very different from that measured at different time of year (Huston 1994).

The greatest threat to biodiversity is habitat loss, degradation, and fragmentation.  Altered habitats do not necessarily eliminate all species, but change heterogeneity, disturbance dynamics, and resources so that some species are reduced or eliminated, and other species become more abundant (Szaro 1996).

The invasion and spread of nonnative species is the second leading threat to genetic and species diversity on wildlands.  In the southern Appalachians, exotic plants, insects, and diseases have dramatically altered forest landscapes.  Chestnut blight, for example, wiped out American Chestnut over its entire range.  Humans facilitate the invasion of non-native species in two ways.  First, by introducing organisms to foreign environments, either intentionally or unintentionally.  Second, by making natural ecosystems more susceptible to invasion by disturbing plant communities through plowing, mowing, harvesting, grazing, and trampling (Szaro 1996).

Other threats to species diversity include air and water pollution, overharvesting of both timber and nontimber forest products, climate change, and uncontrolled fires.

Ecosystem resilience, one of several measures of ecosystem stability, is the degree, manner, and rate of recovery of an ecosystem to a pre disturbance condition (Majer 1989). Although debated, biodiversity is thought to increase ecosystem resilience. One argument is that greater species diversity provides more pathways for recovery. For this reason, some proponents of sustainable forestry require that forest managers maintain both species and structural diversity in managed stands. Often, this objective is approached by managing structural diversity directly, thereby increasing species diversity indirectly. This approach argues that forests that have a diverse vertical structure will generally support a greater variety of animal life than forests with a simple structure: the greater the number and spatial variability of canopy layers and canopy tree species, the greater the possibility for specialization in the animal community and, in theory, the greater the number of animal species. Increasing structural diversity involves the development of a multilayered tree canopy, the retention of a minimum number of large rotting logs on the ground, and the maintenance of large standing dead tree stems in the stand (Kimmons 1997).

Kimmons (1997) points out that, however, the generalization that more diverse forests are more stable forests is not always supported by data. In some forests, some measures of stability will increase as various measures of biodiversity increase. In other forests, there appears to be no simple relationship between various measures of diversity and stability (Kimmons 1997).

Actively managing for biodiversity is a relatively new idea.  Previous efforts to conserve biodiversity focused almost exclusively on habitat preservation: excluding management activities from designated areas.  Even today, managing species diversity is rarely a primary management objective, except where mandated by the Endangered Species Act.  Yet, virtually all management actions, including the decision not to take action, can affect biodiversity at several scales.  Therefore, forest managers need to be aware of both the positive and negative impacts of management activities on biodiversity.  In the southern Appalachians, land managed for timber production constitutes 95 percent of total forested land (SAMAB 1996).  Given both the high levels of biodiversity of this region and the importance of timber production in local economies, the impacts of forest management on biodiversity is an important issue for forest managers and conservationists alike.  From a biodiversity maintenance perspective, forest management for timber is preferable to most other land-use practices (such as agriculture and urban development).  Therefore, managed forests can play a positive role in the conservation of biodiversity.  On the other hand, timber management can have potentially negative impacts on biodiversity.  Although a comprehensive analysis of the effects of timber management on biodiversity in the southern Appalachians have not been analyzed comprehensively, a report by Putz and others (2000) summarizing the impacts of timber management on landscape, ecosystem, community, species, and genetic diversity in tropical forests provides a useful framework for assessing these effects. A graphical model of these impacts of timber management in tropical forests may also be applicable to southern Appalachian forests (Putz and others 2000).

Conceptual Model of the Impacts of
Forest Management on Biodiversity

• Virtually all management actions, including the decision not to take action, can affect biodiversity (Szaro 1996).
• More is not necessarily better. Introducing disturbance or exotic species to an area may increase species richness, but often at the expense of rare species. Although decisions about which species or community types are most valuable are not straightforward, rarity of species should always be considered (Putz and others 2000).
• Sustainable natural resource management will help to conserve most components of biodiversity.  Significant declines in species diversity are often a signal that natural resources are not being managed well (Szaro 1996).
• Focus on function.  Management objectives may be more effectively directed at maintaining functional components of biodiversity rather than at maintaining the total number of species. However, given that the functional properties of many species are unknown, it should never be assumed that a particular species is redundant or unimportant (Szaro 1996).
• Proper land use planning can reduce conflicts. There is no inherent conflict between biodiversity conservation and sustainable resource use. One reason is that the diversity of many organisms tends to be highest in marginally productive land that is not ideal for intensive human land use.  Land use planning can assure that each part of the landscape is used for those purposes for which it is best suited (Szaro 1996: 17).
• Impacts of management activities should be evaluated in relation to the natural influences on species diversity. A specific management action, such as adding fertilizer or water, can either increase or decrease species diversity, but the ultimate outcome will be influenced by natural conditions, such as a local drought. Therefore, knowledge of local conditions and how they may be changing is important for anticipating the impacts of management actions (Szaro 1996).

A number of laws have been passed that focus on the protection of individual species. The most important these laws is the Endangered Species Act passed in 1973. The Endangered Species Act of 1973 requires all Federal agencies to undertake programs for the conservation of endangered and threatened species, and prohibits them from authorizing, funding, or carrying out any action that would jeopardize a listed species or destroy or modify its "critical habitat".  The U.S. Fish and Wildlife Service and the National Marine Fisheries Service became the designated Federal agencies with the responsibility for administering the law.

The 1973 act defines an endangered species as "any species which is in danger of extinction throughout all or a significant portion of its range" and threatened species as "any species which is likely to become an endangered species within the foreseeable future throughout all or a significant portion of its range" (Endangered Species Act 1973, Sec. 3).

More than 950 species of plants and animals native to the United States, and more than 560 species living in other parts of the world, have been placed on the U.S. List of Endangered and Threatened Wildlife and Plants, and thus receive protection under the Endangered Species Act. In the southern Appalachians, 17 animal and 34 plant species have been declared either threatened or endangered.

Legislation Leading to the Endangered Species Act

The U.S. Congress passed the Endangered Species Preservation Act in 1966.  That act allowed listing of native animal species as endangered and provided some means for their protection. The Endangered Species Conservation Act of 1969 was passed to provide additional protection to plants and animals in danger of "worldwide extinction." The Convention on International Trade in Endangered Species of Wild Fauna and Flora (CITES), signed in 1973, restricted international commerce in plant and animal species and their parts and products where such commerce was harmful to the species.

See also: U.S. Fish and Wildlife Services full-text version of the Endangered Species Act

See also: Enviro-link Networks online resource guide to the Endangered Species Act

Protected areas are essential for biodiversity conservation, but many authors have argued that protected areas are not enough to assure the continued existence of the majority of natural landscapes, ecosystems, communities, species and genotypes (Flamm 1988, Zahner 1996, Putz and others 2000). For example, in the southern Appalachians, roadless and wilderness areas constitute only 4 percent of total land area (SAMAB 1996). Consequently, priority must also be given to the conservation of biodiversity outside of these protected areas.

In the southern Appalachians, land managed for timber production constitutes 95 percent of forested land (SAMAB 1996) and therefore presents great potential for maintaining biodiversity. From a biodiversity maintenance perspective, forest management for timber is preferable to most other land-use practices other than complete protection. The number of taxa with reportedly inconsistent, neutral, or positive responses to logging suggests that timber harvesting is not necessarily incompatible with the maintenance of biodiversity.

The deleterious effects of forest management practices on biodiversity are potentially great if harvesting methods are unplanned and uncontrolled. Yet these impacts can be greatly curtailed by adhering to reduced impact logging guidelines such as BMPs (Best Management Practices). Several alternative silvicultural practices, such as leaving ecologically important areas intact within managed areas, can also increase the contribution of managed forests to the conservation of biodiversity conservation. Putz and others (2000) admit that managed forests will not replace protected areas as storehouses of biodiversity, but they conclude that managed forests can supplement and effectively extend the conservation estate.

Although the landscape component of biodiversity is the least sensitive to logging, changes in the size, spatial distribution, and connectivity of habitat patches across the landscape occur especially as the intensity of management interventions increases. These changes in the habitat mosaic alter species distribution patterns, forest turnover rates, and hydrologic processes (Putz and others 2000).

Structure: Logging may change the size, spatial distribution, juxtaposition, and connectivity of forest patches across the landscape. The degree of isolation of the resulting forest fragments may differ among species. For example, wide logging roads may represent uncrossable barriers for some forest interior species, but secondary vegetation along these roads may be attractive to other species such as deer. Long-term species maintenance in a landscape fragmented by logging is influenced by whether logged stands are interspersed within species-rich forest or in a low diversity landscape such as plantations (Putz and others 2000).

Composition: Logging may alter the identity, distribution, and proportion of habitat types in forests in several ways. Management activities may intentionally create new habitats, such as forests converted into plantations. Application of the same stand "improvement" treatments across the landscape may sacrifice interstand diversity. Logging roads may indirectly foster post-logging habitat changes by creating a conduit for weeds and wildfires (Putz and others 2000).

Function: Logging can greatly influence the permanence of the forest fragments they create by altering landscape-level disturbance regimes. Secondary vegetation along roads, skid trails, and harvesting gaps can increase susceptibility to fire. High intensity and widespread logging may influence hydrological processes. Landscape-level movements of animals can be disrupted by wide logging roads and intensely logged areas. The impacts of roads, however, change with time as the roads and their margins mature (Putz and others 2000).

Implications: Setting aside reserves within logging areas may mitigate some of the negative impacts of logging and other silvicultural treatments.  The full range of landscape features and habitats should be represented within these areas. Small unlogged patches in logged forest can serve as source populations for some species. The degree of fragmentation can be decreased if logging is concentrated in small areas (Putz and others 2000).

The ecosystem component of biodiversity is more sensitive to logging impacts than the landscape component, possibly because management activities are implemented at the ecosystem scale. In contrast to the landscape component, most ecosystem-level impacts are a direct consequence of logging activities.

Structure: Logging may change the biophysical properties of soils and the spatial heterogeneity of forest stands. The degree of change in these structural attributes depends on logging intensity, yarding system, and the care with which the operations are conducted. Soil compaction is most severe with ground-based yarding operations, especially if conducted during wet weather. Mineral soils may be exposed after litter layers and root mats are bladed off by bulldozers. Increased erosion from exposed mineral soils can cause deposition of large volumes of unconsolidated sediments into streams. Road construction and timber removal change the horizontal and vertical complexity of forests. Logging also decreases standing biomass by removing timber and increasing residual tree mortality. Logging also increases stores of necromass, including coarse woody debris. Wildfires and prescribed burns can have obvious effects on biomass and necromass stocks (Putz and others 2000).

Composition: Logging affects ecosystem-level composition by changing biogeochemical stocks. Surface runoff is increased by soil compaction. The limited storage capacity in natural streams can be reduced by sedimentation. Where logging intensity is high and forest recovery is slowed by soil damage, fire, or weed infestations, standing stocks of nutrients in biomass are greatly reduced. Prescribed burns can result in volatilization of nitrogen and sulfur (Putz and others 2000).

Function: Logging impacts ecosystem function by affecting hydrological and biogeochemical fluxes as well as productivity. Logging induced soil compaction may reduce water infiltration into soil and increase surface flow rates and volumes. Soil compaction may also impede root growth, and consequently reduce plant productivity. Disturbance of surface soils may alter nutrient cycling. Sedimentation can greatly modify stream characteristics such as the rations among pools, riffles, and runs ratios. In areas of intensive logging, nutrient cycling and hydrological functions may be greatly modified by reduced canopy interception of rain and mist, decreased uptake of water and nutrients by the diminished biomass, and increased surface erosion. Changes in carbon storage and flux associated with logging and other silvicultural activities influence whether forests are net sources or sinks of "greenhouse" gases. For example, logging-induced transfers of living trees to coarse woody debris can effect understory structure and dynamics leading to more carbon release. Other silvicultural treatments such as prescribed burning, weed control, thinning, and enrichment planting have various impacts on both greenhouse gas emissions and biodiversity (Putz and others 2000).

Implications: Methods for mitigating the ecosystem-level impacts of logging on forests are well known. Alternative yarding techniques such as skyline yarding can reduce damage to residual stands, soils, and streams. Adhering to Best Management Practices (BMPs) can also greatly reduce soil compaction and erosion (Putz and others 2000).

Logging, especially if followed by treatments that increase growth and relative densities of commercial timber species, can substantially change the physiognomy, composition, and trophic structure of forest stands (See: definition of community diversity).

Structure: Logging can change the proportions of forest in mature, recovering, and early successional stages. Silvicultural treatments such as thinning and vine-cutting can increase dominance of commercial species (Putz and others 2000).

Composition: Logging affects community composition by changing the relative abundance of species and guilds inhabiting forest stands. Relative abundances of tree species with light demanding vs. shade tolerant regeneration, wind vs. animal dispersed seeds, vertebrate vs. invertebrate pollinated flowers, and thick vs. thin bark, for example, are all subject to change managed forests. These changes, in turn, alter the composition of different feeding guilds of animals (such as understory insectivores and arboreal folivores) (Putz and others 2000).

Function: Numerous community-level processes, such as pollination, herbivory, seed dispersal and predation, may be affected by logging. Changes in these processes are primarily a consequence of altered resource abundance, which results from logging-induced changes in community structure and composition (Putz and others 2000).

The species component of biodiversity has received the most attention from researchers concerned about the impacts of logging and other silvicultural treatments in tropical forests. Although the most obvious species-level impact of logging is a reduction in the abundance of commercial tree species, the reproduction, growth, and survival of a great number of species can be adversely affected, depending on the intensity of logging.

Structure: The most immediate impacts of logging at the species level are suffered by the harvested tree species whose populations can be greatly depleted, especially in the larger size classes of reproductive individuals. By altering microclimates, logging may affect population structures of specialist species of the forest interior, such as vernal herbs and salamanders (Meier and others 1996). Generalist species with wide-ranging habitats may benefit from logging, as habitat heterogeneity may be increased (Putz and others 2000).

Composition: Although logging decreases density of harvested species, later management activities are usually aimed to regenerate of targeted commercial species and assure their dominance in the new stand. Non-commerical species are also directly impacted by management activities, especially if they are considered to be competitors of commercial species. Populations of other species may be indirectly affected by logging and later management activities, but changes in species composition in response to forestry operations are by no means consistent across or within taxa (Putz and others 2000).

Function: Logging affects key demographic processes such as survivorship, fertility, and recruitment of both the targeted timber species and other species. Logging impacts on tree populations continue for many years after logging is completed because damaged trees suffer high mortality rates, weeds and vines proliferate, interfering with tree reproduction and survival, and population size reduction and fragmentation can decrease pollination levels and change patterns and intensities of seed dispersal and predation. Direct impacts of logging, such as microclimate changes and forest fragmentation, can alter demographic processes of species dependent on mature forests, such as vernal herbs and salamanders (Meier and others 1996).  Logging favors proliferation of disturbance-adapted taxa, some of which are not native or were not previously common in the area.  Thus impacts on the resident flora and fauna can be large (Putz and others 2000).

Implications: Small fragments of unlogged forest can serve as important refuges for plants and animals during and shortly after harvesting. They can serve as sources for colonizers after logging. However, private landowners with small acreages to manage may be unwilling to set aside 10 percent of their forest for species preservation. Therefore, species impacts of logging are of particular concern in small forest management units (Putz and others 2000).

The genetic component of biodiversity is likely to be the most sensitive of all components to logging because of reductions in effective population size and interruptions in gene flow. However, little is known about the impacts of logging on genetic structures because techniques for assessing genetic structure of populations are sophisticated and expensive.

Structure: Ttimber harvesting, silvicultural treatments, forest fragmentation, weed proliferation, and wildfires may all affect genetic diversity by reducing effective population sizes and heterozygosity. Allelic frequencies of commercial species change after removal of a large proportion of healthy reproductive adults, but these changes may be buffered by the presence of advanced regeneration or coppicing. Silvicultural treatments also may affect the genetic structure of species targeted for removal such as woody vines (Putz and others 2000).

Composition: Allelic frequencies may continue to change after timber harvesting due to decreased effective population sizes. Theory suggests that deleterious recessive genes may become more apparent due to dysgenic selection, and heterozygosity may decline due to the "bottleneck" effect in the small, isolated populations (Putz and others 2000).

Function: Logging may affect the functional attribute of the genetic component of biodiversity by interrupting gene flow, which in turn influences outbreeding rates. Decreased effective population sizes coupled with losses of pollinators and seed dispersal agents can result in reduced gene flow and inbreeding depression in populations of both commercial and noncommercial species. However, for species that are dioecious or obligate outcrossers, only very severe reductions in effective population size are likely to have much effect on gene flow (Putz and others 2000).

The term "aesthetics" refers to the visual impact of a landscape. It is often referred to simply as "scenic beauty." It is a forest resource that differs for all people, and is highly valuable in recreational and spiritual terms, but extremely difficult to quantify in monetary terms.

Research has indicated that forest scenes are rated high in aesthetic quality when they contained large trees, low to moderate stand densities, grass and herbaceous cover and a variety of species and colors. Perceptions of a forests scenic beauty are negatively impacted by small trunks, dense shrubs, bare ground, woody debris, and evidence of fire or human disturbance (Ribe 1989).

It is important for forest managers and landowners to understand aesthetics because public support of management practices often depends on how management practices affect aesthetic qualities. It is also important to realize that aesthetic preferences often differ among different ethnic groups.

See also:  Aesthetics and Ethnicity.

Natural resource managers will always work under the watchful eye of the public. Therefore, the human dimensions of forest management are as important as ecological ones. Ultimately, the success of any management practice depends upon public support. A forest practice that is environmentally necessary and economically sound, but aesthetically unacceptable to the public, will either be modified until it is accepted or it will be abandoned (Bliss 2000). Managers need to develop a greater understanding of the local publics aesthetic preferences and, more importantly, how management actions influence those aesthetics.

People search for meaning in landscapes and they apply their own values and standards when judging whether they like or dislike the appearance. Extraction of natural resources, and development of facilities and access routes are often met with opposition and legal obstruction because they are visually unappealing. In the past, resource managers have been criticized for inadequate knowledge of the visual consequences of management actions, lack of alternative actions to satisfy public demands, and lack of understanding of public perception, expectations, concerns and demands (Magill 1990).

Communicating management goals, or explaining the long-term effects of such aesthetically offensive practices as clearcutting or prescribed burning may influence approval levels. In the early 1990s, a series of telephone surveys was conducted to measure public opinion in the Midsouth regarding forest practices and policies. The survey indicated that many opponents of clearcutting softened their opposition to the practice, when they were assured that harvested areas would grow back to trees (Bliss and others 1997). Often the level of approval of aesthetically unpleasant practices depends on land ownership. The southern phone survey also indicated that three times as many respondents found the practice of clearcutting acceptable on privately owned land as on public land (Bliss and others 1997).

A Tucson study asked participants to rate slides of forest scenes for scenic quality and acceptability for recreation. The scenes showed ponderosa pine (Pinus ponderosa Laws.) stands that were either unburned or had experienced light or severe fire 1 to 5 years previously. Participants were given brochures about fire effects, and were retested after they read that material. Ratings indicated that scenic quality was improved by light fire but diminished by severe fire. Levels of acceptability differed with the type of recreation considered, with camping showing the greatest sensitivity to fire effects. Brochures increased respondents knowledge about and tolerance of fire but did not affect ratings of scenic recreational quality (Taylor and Daniel 1984).

See also: Effects of Fire on Aesthetics and Using Fire to Enhance Appearance

Research shows that wildland areas are not equally aesthetically pleasing to all groups of Americans (Rideout and Legg 2000, Wallace and Witter 1993, Washburne 1978). Rideout and Legg (2000) conducted a focus-group study to determine outdoor wildland-related recreation preferences and barriers among different ethnic or racial groups in Fort Worth, Texas. American Indians and Caucasians discussed viewing scenery more often, and it ranked more importantly with them, than with African Americans and Hispanics.

This finding is supported by other studies. In 1993, a study of perceived attractiveness of tourism destinations asked African American and Caucasian participants to rank three sets of 10 color photographs from most to least attractive (Philipp 1993). The most significant difference was in the ranking of the picture that the author believed most represented wildland scenes. Caucasians ranked the picture of mountains considerably higher in attractiveness than African Americans did (Philipp 1993).