Literature Review Part B: Ecological Aspects of Naturalness

The recorded study of what we now call ecology dates from at least the time of Theophrastus, although the term "ecology." or, "oekologie" was first formally proposed by Ernst Haeckel in 1869 (Barbour, Burk, and Pitts 1987). Over time, ecology came to encompass the study of biotic communities and broadly conceived relationships between their component parts, especially through the contributions of Clements, Shelford, and others in the 1920s (Odum, E.P. 1983). In 1935, A.G. Tansley proposed the term "ecosystem," which today is widely regarded as a central concept in our understanding of ecology (Cherrett 1989), and which has become a focus of much research, to the point that we now speak of "ecosystems ecology" as a field of study in itself (Odum, E.P. 1983).

In recent decades, moreover, a fundamental change has occurred in the ecological sciences, as what was once chiefly "field science," or observation, has become more applied. Today, we no longer simply study ecosystems, but seek to actively and experimentally manage them, for a variety of extractive and environmental management purposes (Bormann et al. 1994; Pomeroy, Hargrove, and Alberts 1988).

Over time, ecology has developed into an integral part of most aspects of resource management. Forestry is no longer a simple matter of cutting trees, but instead an ecologically based mixture of silviculture and other biological disciplines (O'Hara and Oliver 1992). Likewise, agriculture (Jackson 1991), grazing (Savory 1988)--even such superficially non-biological forms of resource management as mining (Thornburg 1982) or industrialized land use planning (Van Riet and Cooks 1990)--have become linked with an

ecological framework. This linkage can be seen in the passage of the National Environmental Policy Act, the Endangered Species Act, the Clean Water Act, and many other similar laws, and the resultant requirements for preparation of Environmental Assessments or Environmental Impact Statements. Whether one is strip mining coal or planting pine trees, it is a given today that management of natural resources requires planning and processes grounded in ecology.

Much of the past application of ecological theory to resource management, however, can be seen as more reactive than proactive. Risk assessment (Suter 1993; Wilson 1991), environmental impact assessment (Burchell and Listokin 1975; Shipley 1990), and environmental monitoring (Clarke 1986; Goldsmith 1991) all play important roles in policy development and decision making, but their focus is more upon mitigation of problems associated with resource use, rather than the activities of on-going management or planning and implementation at a project level.

As a recent and still-emerging approach to managing natural resources, "Ecosystem Management" seems particularly promising as a framework for incorporating the concepts of ecology directly into active management. Unlike risk or impact assessment, this approach comes to resource management asking how ecological theory can be used to achieve goals without surpassing ecological constraints, instead of simply seeking to avoid inflicting ecological damage in the pursuit of such goals. The Ecosystem Management framework takes this more pro-active approach by focusing on an understanding of dynamic ecological structures and processes, especially by seeking an understanding of ecosystems and landscapes in terms of their natural variability, natural disturbance regimes, and other natural processes.

Given this context, an examination of the Ecosystem Management framework serves to highlight the ways in which ecological principles and techniques may be incorporated within a management framework--and it also serves to highlight the ways in which the concept of naturalness underlies such management as well.

Ecosystem Management is a relatively new approach to natural resource management. Ecosystem Management ("EM") has developed as an outgrowth of other proposed management frameworks, such as "Adaptive Management" (Holling 1976; Walters 1986), "New Perspectives" (Franklin 1992; McQuillan 1992), and "New Forestry," (Agee 1993; Maser 1992), which can be together seen as an overall shift in management philosophy, often heralded as a "paradigm shift" in North American forestry (Kaufmann et al. 1994; McConnell et al. 1994; Thomas 1993).

Ecosystem Management has been promoted in large part because of a perception that past management actions failed to perform well ecologically (Goldstein 1992; Pell 1994), as well as a perceived need for better integration of ecological and societal values in planning management prescriptions (McConnell et al. 1994; Pell 1994). In essence, these concerns have resulted in recognition of the need for better objective, scientific input to the political decision making processes of resource management (Iverson 1993; Thomas 1993).

Building upon the theoretical contributions of various scientists, such as Shelford, Tansley, Lindeman, and Odum, who decades earlier had promoted both dynamic and systems-based approaches to ecology (Golley 1993; McIntosh 1985), numerous ecologists in the 1970s and 1980s began to promote an ecosystem-based approach to land management (Grumbine 1994), calling for a change from ecology as a theoretical science to an "applied ecology" which would become "a preventative and regulatory concern" (McHale 1970).

References specific to "ecosystem management" as an approach to resource management appear in the literature primarily within the past fifteen years, (Francis et al. 1979; Magnuson et al. 1980), but the underlying concepts have been traced to the 1930s and 1940s, the same time period in which the American conservation movement was becoming active (Grumbine 1994), and the concept of the "ecosystem" was itself first being introduced and elaborated (Golley 1993).

The groundwork for what might now fall under the EM heading was laid by landscape architect and land-use planner Ian McHarg, who in the mid-1960's advocated an ecologically based approach to land-use. McHarg promoted a blending of ecological science techniques with principles of landscape-based studies, emphasizing the need for "ecosystem inventory" and assessments of ecosystem "health" in jointly managing the natural and cultural aspects of landscapes (Belknap et al. 1967; McHarg 1969).

This kind of approach to resource management was further promoted through the idea of biological conservation. In the inaugural issue of the journal Biological Conservation, J.B. Cragg (1968), suggested a definition of biological conservation as being: "concerned with the maintenance of natural ecosystems and, where possible, their utilization--either directly or by way of the information obtained from their study--for the long-term benefit of mankind." Nearly thirty years later, this stands as remarkably similar to most attempts at defining an EM approach to resource management.

Another important step toward developing an EM approach came in the work of C.S. Holling (1976) who called for resource managers to move beyond a "reactive" stance and to actively employ ecological knowledge in improving their management efforts through an "adaptive" approach. This idea of "adaptive management" was then further developed into a specific framework for resource management, which sought to employ experimental methods of ecology while acknowledging that a degree of uncertainty was inevitable in attempting to manage complex ecological systems (Salwasser 1986; Walters 1986). Such an adaptive, or experimental approach to management, can be seen as a central tenet of the emerging EM management framework (Grumbine 1994).

Development of EM into a specific management strategy has been particularly active within the U.S. Forest Service, which views EM as an opportunity for refinement of the agency's earlier "Multiple Use Sustained Yield" approach to resources (Jensen and Everett 1994; Salwasser et al. 1993). In 1992, Ecosystem Management was officially adopted by the U.S. Forest Service as that land management agency's approach to sustainable resource management, thus establishing EM as more than a merely theoretical framework (McConnell et al. 1994).

Much of the extant literature regarding EM has in fact been produced by the U.S. Forest Service, and the influence of this agency on EM's development must be recognized; however, it is also important to acknowledge that many concepts incorporated into EM come from earlier or parallel planning models, some of which focus on wildlife, soil science, agriculture, and engineering rather than on forestry per se (Grumbine 1994; Jensen and Everett 1994).

Additionally, other approaches, although not identical to "Ecosystem Management," exist as parallel efforts. These include: "New Forestry," promoted chiefly in the Pacific Northwest region of the United States (Agee 1993; Franklin 1993; Maser 1992); "Natural Forest Management," promoted for tropical rain forest management (Gorchov 1994); the "Richer Forest" program of Sweden (Salwasser, MacCleery, and Snellgrove 1993); and "Close-to-Nature" management, practiced in parts of Central Europe (McConnell and Diaci 1996).

More specifically, "EM" itself is not limited to the U.S. Forest Service. Other government agencies (Grumbine 1994), other governments outside of the United States, (Slocombe 1993), and non-governmental groups interested in conservation (Goldstein 1992; Grumbine 1994) have all recently embraced an EM approach. As a result of these varying origins and perspectives, varying interpretations of EM certainly exist (Grumbine 1994).

While much of the current literature on EM seems to indicate a fairly homogeneous view of what EM means, it does bear emphasizing that controversy and disagreements exist, and that some have come to regard the entire concept of EM with a measure of suspicion (Donnelly 1995; Stanley 1995). Against this background, it is essential to recognize that, while a more or less general picture can be formed of what EM is, all of these diverse perspectives do have the potential to shape the future of EM as it continues to develop.

EM was developed as a system or process for resource management (Agee 1993; Bormann et al. 1994; Grumbine 1994; Scientific Integration Team 1994), not as a more general approach to ecology or science. However, EM is also recognized as an approach to management which is grounded in ecological principles and knowledge of ecological processes (Bourgeron et al. 1994; Kaufmann et al. 1994).

Ecology, however, is itself a broadly defined science with many distinct theoretical approaches and sub-disciplines (Golley 1993; McIntosh 1985), and as such, it is important to consider which aspects of ecological science are specifically emphasized within an EM approach. The most prominent of these are the perspectives of ecosystem ecology and landscape ecology, both of which are seen as critical and complimentary aspects of sound ecological science (Urban, O'Neill, and Shugart 1987) and resource management (Bourgeron et al. 1994; Franklin and Forman 1987).

Although ecosystem ecology and landscape ecology differ substantially in their respective methods and aims (Forman and Godron 1986; McIntosh 1985; Naveh and Lieberman 1984; Odum 1989; Slocombe 1993), they are often treated as nearly synonymous in the context of EM. For example, concepts from landscape ecology have been described as "ecological principles for ecosystem management" (Bourgeron and Jensen 1994), landscapes have been described as "the context for planning management of ecosystems" (Salwasser et al. 1993), and it has been stated that "landscape ecology principles provide the foundation for ecosystem management" (Jensen, Bourgeron, Hessburg, et al. 1994).

Ecosystem management, as its name implies, is expressly based in the concept of the ecosystem, and hence in ecosystem ecology (Bormann et al. 1994; Grumbine 1994). Although social and economic concerns are the objective focus of EM's management prescriptions, it is recognized that the key to achieving management objectives in a sustainable way is an understanding of ecological functions and capacities (Bormann et al. 1994; Kaufmann et al. 1994) specifically in terms of using natural, or unexploited ecosystems as a comparative baseline against which to evaluate managed systems (Noss 1993). In this regard, EM is dependent upon sound application of the principles and methods of ecosystem ecology.

Likewise, numerous EM applications have incorporated landscape ecology principles and methods as an integral part of their planning process. This includes broadly defined multiple-use forest management planning efforts (McConnell et al. 1994; O'Hara et al. 1994), timber-harvest planning (Franklin and Forman 1987; Mladenoff et al. 1994), and planning related to mitigating the effects of fire suppression (Baker 1994; Shlisky 1994).

Ecosystem ecology refers specifically to a branch of biological science which developed in response to A.G. Tansley's proposed concept of the ecosystem, and the subsequent work of other biologists who then elaborated a systems-based approach to ecology (Odum, E.P. 1983). The ecosystem concept, as developed by the work of Tansley, Lindeman, Whittaker, and Odum, among others (Golley 1993; McIntosh 1985) itself entails several key elements which have particular impact on how the EM framework defines its objectives and methods.

Defined most simply, ecosystem ecology is the study of ecosystems, an approach to ecology concerned with the dynamic ecosystem as its basic unit of study and which seeks to identify and describe ecosystem properties and processes (McIntosh 1985). Ecosystem ecology is therefore based foremost in the practice of observation; observation is in turn based in the recognition of change, and change is based in recognition of structures at given moments in time (Allen and Hoekstra 1992). In this sense, ecosystem ecology is focused on "natural" phenomena, rather than "formal" models (Louie 1985), and it is therefore regarded as more descriptive than analytical at its most fundamental level (McIntosh 1985).

However, because ecosystem ecology seeks to describe complex whole systems which elude simple description, it must also employ abstract models to explicate the complexity of its subject (Odum, H.T. 1983). This two-step approach to ecosystem ecology, observation and description of natural phenomena, matched with predictive modeling of more complex system properties

based upon those observations, can be seen as a key to the "adaptive" approach of EM (Grumbine 1994; Pfister 1993). Only through observation of natural systems and processes will the EM framework be able to gauge the long-term effects of given management actions and devise strategies for sustainable resource use (Kaufmann et al. 1994; Salwasser et al. 1993).

Ecosystem ecology also offers two specific perspectives which are important to the EM approach. First, the concept of the ecosystem, as originally developed by Tansley, offered a mechanistic alternative to the "supraorganism" paradigm held by ecologists of the past century (Evernden 1985; Rich 1988). This mechanistic or systems approach was then elaborated by Lindeman and others into the trophic-dynamic or energetics approach now fundamental to ecology (Golley 1993; McIntosh 1985), which sees trophic flows, or the inputs and outputs of a given system, as a way of grasping the dynamic processes of ecosystems which cannot be inferred from static descriptions of ecosystem structure (Odum, E.P. 1983). As a further extension of this perspective, ecosystem ecologists often employ the "black box" model in describing ecosystems, in which it is recognized that, while ecosystems are too complex to be understood in their entirety, their overall character can nonetheless be inferred from modeling based in observation of their inputs and outputs (Baser 1976; Stark 1966).

Within the context of EM, this emphasis on the dynamic character of ecosystems is reflected in the generally accepted assertion that true ecosystem "equilibrium" or stability is at best rare, or merely theoretical (Golley 1993; Odum 1989), and that system processes or dynamics, rather than simple, static descriptions of system composition, are key to understanding ecosystem characteristics (Bormann et al. 1994; Kaufmann et al. 1994; U.S. Department of Agriculture, Forest Service 1994).

In a more applied sense, the EM literature also incorporates this understanding of the dynamic character of ecosystems by employing concepts such as resilience, persistence, resistance, and variability (National Research Council 1986; Pimm 1991) in the description of system properties, and by focusing on the role of disturbance and variability in maintaining natural ecosystems (Bourgeron et al. 1994; Swanson et al. 1994). EM also incorporates elements of the "black box" model, recognizing that it is simply not practical to attempt management of each individual element of an ecosystem, and that managers must instead focus on understanding ecosystem organization and the effects of management actions at this level (Jones and Lloyd 1993).

A second key implication of the ecosystem ecology perspective is that the unit of the ecosystem is viewed as constituting one scale of observation within a nested hierarchy, ranging from individual atoms up to the entire universe; in this regard ecosystem ecology generally views populations and communities as smaller units encompassed by ecosystems (Odum 1989; Shugart and Urban 1988). Ecosystem ecology thus takes a complex view of ecosystems and applies principles of both system theory (Bertalanffy 1969, 1975) and hierarchy theory (Pattee 1973) to understanding ecological interactions, using observations at lower levels to understand higher-order organizational processes through concepts of feedback loops, nutrient pathways, and causal forces (Odum, E.P. 1983).

Within the EM framework, this hierarchical perspective has been employed in a tendency to focus on "higher-order" levels of ecosystem organization (Odum, E.P. 1983). For example, the EM literature generally places importance on understanding overall ecosystem function rather than specific small-scale processes (Bormann et al. 1994; Franklin 1993; Jensen, Bourgeron, Hessburg, et al. 1994), and ecosystems are often described within broadly defined classification systems (Bourgeron, Humphries, and Jensen 1994; Jones and Lloyd 1993) rather than in terms of site-specifics.

Likewise, EM proponents tend to focus monitoring and assessment on select indicator species (Marcot et al. 1994) and indices of general ecosystem "health" (Comanor 1993; Scientific Integration Team 1994), "integrity" (Grumbine 1994; Iverson and Cornett 1994), or "sustainability" (Kaufmann et al. 1994; Salwasser et al. 1993), rather than on more detailed site analysis. EM is seen in this regard as a management framework which aims to achieve these more broadly defined goals rather than protecting or preserving "natural" conditions (Agee 1993; Grumbine 1994) or micro-managing all individual ecosystem components (Bourgeron et al. 1994; Kaufmann et al. 1994).

This hierarchical approach is also fundamental not only to ecosystem ecology, but to landscape ecology as well (Wiens 1995), and it is through a hierarchical approach that EM can be seen to bridge between other elements of both ecosystem ecology and landscape ecology, working with aspects of each to achieve management objectives (Bourgeron and Jensen 1994; Franklin and Forman 1987; Kaufmann et al. 1994).

Landscape ecology has been defined as the study of "how a heterogeneous combination of ecosystems is structured, functions, and changes" (Forman and Godron 1986), or "the study of landscape which interprets it as supporting natural and cultural systems" (Vink and Davidson 1983). The principles of landscape ecology are frequently mentioned as forming an integral part of the management strategies which EM entails (Grumbine 1994; Jensen, Bourgeron, Everett, et al. 1994; McConnell et al. 1994; Slocombe 1993).

Landscape ecology's development from roots in geography occurred roughly within the same time period as ecosystem ecology's development as a branch of biology. The origin of landscape ecology has been traced to the work of F.H.A. von Humboldt (Allen and Hoekstra 1992), who proposed the concept of ecological "associations" as a descriptive unit and developed plant geography as a science in the 1800s (Barbour, Burk, and Pitts 1987). What we now call landscape ecology, however, may perhaps be most closely attributed to the work of Siegfried Passarge, who in 1919 proposed a system of analysis he termed Beschreibende Landschaftskunde, or "Descriptive Landscape-science," defined as "systematic observation of the phenomena which compose the landscape" (Sauer 1963).

The term "landscape ecology" is itself credited to Carl Troll, who in the late 1930s elaborated upon the ideas of Passarge and others with a perspective based in the analysis of aerial photography (Forman and Godron 1986; Golley 1993), and whose own work was further developed by various geographers and land-use planners working in Germany, the Netherlands, and the Soviet Union during

the 1950s and 1960s (Golley 1994; Wiens et al. 1993). In the 1980s, landscape ecology gained a measure of acceptance in the United States, with the successive publication of three influential texts: Landscape Ecology and Land Use, by A.P.A. Vink and D.A. Davidson in 1983; Landscape Ecology: Theory and Application, by Z. Naveh and A.S. Lieberman in 1984; and Landscape Ecology, by R.T.T. Forman and M. Godron in 1986.

Landscape ecology is seen as resting on three key principles: (1) the concept of landscapes as spatially heterogeneous areas composed of identifiably homogeneous elements; (2) the idea that fluxes and redistributions of materials and energy occur among these component landscape elements; and (3) the idea that the landscape itself is shaped by both human cultural actions and natural environmental processes, which in turn also have reciprocal influences upon each other (Risser 1987). Perhaps most importantly, the perspective of landscape ecology is seen as one based in a central concern with the concept of landscape heterogeneity (Forman and Godron 1986; Risser 1987), and, derived from this concern, attention to issues of scale and the application of hierarchy theory as an approach to scale (Wiens et al. 1993; Turner et al. 1993).

Landscape ecology has been described as being primarily concerned with discerning patterns from which underlying causative mechanisms may be inferred (Risser 1987; Wiens 1995), and because of this focus landscape ecology has developed terminology which allows for the systematic description of landscape components. Elements of landscape structure are categorized as "patches," the basic homogeneously identifiable units of landscapes; "corridors," narrow features of connectivity between patches; "networks," aggregates of corridors; and "matrix," larger landscape units within which these other features are embedded. Each of these elements is in turn described using a variety of terms, such as "extent" and "shape," and "porosity" to clarify their origins and delineate their spatial relationships. The overall structure of landscapes is in turn viewed as heterogeneous "configurations" or "mosaics" composed of these elements and connected functionally by ecological dynamics (Forman and Godron 1986; Wiens et al. 1993).

Elements of landscape function consist of both "natural" processes, including geomorphologic, edaphic, evolutionary, and successional processes, and "human" processes, including disturbance of natural processes on various spatial and temporal scales, landscape modification, and development or urbanization; "flows" between landscape elements, including atmospheric, hydrological, and erosion processes; and "animal and plant" processes, including migration and dispersal (Forman and Godron 1986). These are for the most part further described within a landscape ecology context simply by using appropriate terminology borrowed from the ecological or social sciences.

Elements of landscape change include measures of "stability," instability," "patterns of change," and "landscape dynamics." These are in turn described with terms borrowed from ecosystem ecology, such as "variability," "persistence," "resistance," and so forth, to describe the spatial and temporal characteristics of the processes and their influence upon the landscape (Forman and Godron 1986).

In a general sense, landscape ecology is invoked within the EM framework through an emphasis placed on pattern and process, or structure and function, as the basic elements of landscapes amenable to study for planning purposes (Jensen, Bourgeron, Hessburg, et al. 1994; Kaufmann et al. 1994; Scientific Integration Team 1994). Inference of process from pattern is considered the basic tool of landscape ecology, and has been described as "key to the development of principles for land management" within the EM framework as well (Bourgeron and Jensen 1994).

More specifically, EM applications have made use of landscape ecology terminology for describing landscape components, especially the ideas of "patch," "matrix," and "mosaic landscapes" (Franklin 1993; Jensen, Bourgeron, Everett, et al. 1994; Shlisky 1994; U.S. Department of Agriculture/U.S. Department of Interior 1994). The use of these terms has in turn been employed in addressing specific management concerns such as defining the natural range of variability (Bailey et al. 1994; Swanson et al. 1994) defining ecosystem integrity and health (Bourgeron et al. 1994; King 1993), analyzing landscapes for habitat fragmentation and connectivity (Noss 1987; Turner et al. 1994), and planning for the preservation of biological diversity (Franklin 1993; Noss 1990).

Perhaps the most specific way in which landscape ecology principles have been incorporated within the EM framework is in regard to issues of scale and hierarchy. Application of hierarchy theory, as employed by landscape ecology, is seen as a basic approach to delineating analysis areas and performing assessments at various spatial scales (Hann, Jensen, et al. 1994; Kaufmann et al. 1994; McConnell et al. 1994; Scientific Integration Team 1994). In particular, the U.S. Forest Service has invoked landscape ecology concepts of hierarchy theory in placing a strong emphasis on analysis at the scale-level of watersheds (Bormann et al. 1994; Kaufmann et al. 1994; U.S. Department of Agriculture, Forest Service 1994b), which are regarded in landscape ecology as sub-units of landscapes (Forman and Godron 1986).

The EM literature appears to have also borrowed from landscape ecology in two other important matters: in developing models for planning and implementation, and in incorporating social or cultural elements along with ecological ones into its analysis and decision making component. Unlike ecosystem ecology, landscape ecology places an equal emphasis on the cultural and natural aspects of the landscape (Forman and Godron 1986; Thorne and Huang 1991; Vink and Davidson 1983), and this has been significant in enabling application of landscape ecology to the development of EM as a management framework, as opposed to merely contributing data upon which management may be based.

Even from a glance at their schematics, planning models proposed for EM (Bormann et al. 1994; Oliver, Knapp, and Everett 1994; Scientific Integration Team 1994) bear strong resemblance to models previously proposed within the framework of landscape ecology (Steiner and Osterman 1988; Van Riet and Cooks 1990; Vink and Davidson 1983). As with landscape ecology, planning within the EM framework is seen as attempting to define linkages across scales through hierarchical analysis, and attempts to blend social and ecological aspects of planning within the decision making process (Bormann et al. 1994; Oliver, Knapp and Everett 1994; Scientific Integration Team 1994).

Additionally, these EM planning models also make explicit reference to landscape ecology in the context of their focus on humans as components of ecosystems (Golley 1994), and proposals for implementing an EM strategy include efforts to "place society squarely within ecosystem management" (Bormann et al. 1994). In this regard, EM has been described as a two-way relationship of elements, meshing human social and economic wants with the bio-physical constraints of ecosystem capacities (Bormann et al. 1994), or as a three-way relationship between social human desires, technological/economic factors, and ecological capabilities (Jensen, Bourgeron, Hessburg, et al. 1994). The ideas of "humans embedded in nature" (Grumbine 1994), and of landscapes being shaped more or less equally by both natural and cultural processes, are central to the EM framework (Kaufmann et al. 1994; Oliver, Irwin, and Knapp 1994; Scientific Integration Team 1994), and to landscape ecology as well (Forman and Godron 1986; Naveh and Lieberman 1984; Vink and Davidson 1983).

Drawing upon diverse concepts borrowed from the literature of ecosystem ecology, landscape ecology, biological conservation, and adaptive management, proponents of an EM framework for natural resource management have developed various lists of key principles or methodologies. It is perhaps indicative of the as-yet developing character of EM that there are nearly as many lists of guiding principles as there are proposals for implementing an EM approach; nevertheless, some themes seem common to a majority of these proposals.

Procedurally, it has been suggested that there are three fundamental steps to implementing an EM approach: designating the physical boundaries of the system of interest, understanding the interactions between the parts that function as a whole system, and understanding the relation between the system and its context (Bormann et al. 1994). This set of guidelines can be seen as reinforcing the basic ecosystem and systems-theory aspects of EM, and also serves as a frame for understanding the importance variously placed on other key concepts or principles--all of which can be seen as contributing to these three steps of analysis.

Perhaps the most important point raised by various proponents of an EM framework is an emphasis on the idea that EM, as a management system, represents an approach to satisfying social and economic pressures (Bormann et al. 1994). In this respect, the key to EM is portrayed as the need to balance the social and human elements of management against the ecological constraints of the systems being managed (Grumbine 1994; Human Dimension Study Group 1994; Pell 1994; Sample et al. 1993).

This acknowledgment of the primary orientation of EM toward meeting social pressures has also resulted in the view that EM must be based in a cooperative or participatory approach, allowing management's constituencies ample input to decision making and transcending narrow definitions of responsibility (Grumbine 1994; Human Dimension Study Group 1994; McConnell et al. 1994; Slocombe 1994). Thus, EM, while oriented toward principles of ecological science in the procedures of its analysis elements, remains oriented toward the production of "goods and services" for society in its decision making and resource allocation steps.

These combined social and ecological aspects of EM are also seen as requiring the grounding of EM implementation in elements of an ecosystem ecology perspective, recognized chiefly in the literature by an emphasis on ecosystems as dynamic and variable (Grumbine 1994; Kaufmann et al. 1994; Pfister 1993; Scientific Integration Team 1994). The importance of such ecological concepts is also matched with an emphasis on elements of the landscape ecology perspective, especially the elements of a hierarchical perspective and attendant concepts of scale (Grumbine 1994; Kaufmann et al. 1994; Pell 1994; Scientific Integration Team 1994). While socially derived goals may drive the processes of EM, these ecosystem and landscape ecology principles are seen as the basic tools of EM research and implementation.

The need to balance the human and ecological aspects of ecosystem-based management is in turn variously expressed by attention to ecological integrity (Grumbine 1994; Pell 1994), the maintenance of ecological resiliency (Kaufmann et al. 1994), the

notion of ecological limits (Scientific Integration Team 1994), or concepts of ecological sustainability (Human Dimension Study Group 1994; Kaufmann et al. 1994; McConnell et al. 1994; Salwasser et al. 1993). And these concerns in turn suggest the need for a flexible, experimental, or "adaptive" approach to management, recognizing that the natural variability of ecosystems makes long-term prescriptions difficult to assess without on-going monitoring and adjustment of methods (Grumbine 1994; Pell 1994; Scientific Integration Team 1994).

As discussed above, some see EM primarily as a way to better incorporate ecological understanding within an otherwise politically and socio-economically based decision making process (Goldstein 1992; Kaufmann et al. 1994). Conversely, others see EM as a way of giving a more clearly primary role to social wants or "desires" in resource management decisions (Bormann et al. 1994), to the point where it may be argued that all issues in ecosystem management must be thought of as being "social" by nature (Allen 1995). In this latter view, ecosystems are seen as "patterns and processes required to produce products" for human society (Bormann et al. 1994). Interpreted in this way, EM can be seen primarily as a resource management approach, as opposed to a subset of ecological science, although this is perhaps not a readily apparent, nor an apparently necessary, approach to all parties concerned (Donnelly 1995).

To at least some extent, this has resulted in two "camps" regarding EM; those who place emphasis on the primacy of "ecosystem" and those who place emphasis on the primacy of "management" when discussing the essence of EM (Donnelly 1995). These can also be differentiated as fundamentally biocentric versus anthropocentric (Stanley 1995), or ecological versus political (Baines 1989) approaches, perhaps reflecting the basic divergence between what have been described as the "nature-endorsing" and "nature-skeptical" philosophies toward the natural world (Soper 1995).

Although most of the literature specific to EM has been written from a resource management perspective, it is apparent that these differing emphases each have their adherents, a fact which should be borne in mind when considering the definitions and implementation strategies for EM proposed by various authors. In part because of these differing philosophical perspectives, and perhaps also in part because the concept of Ecosystem Management is yet in its development, there are a number of proposed definitions of EM in circulation among resource managers.

"Ecosystem management involves regulating internal ecosystem structure and function, plus inputs and outputs, to achieve socially desirable conditions. It is a process of understanding ecosystem components (including people) and interactions" (Agee 1993).

Although these definitions are quite general, they point to the twin focal points of the EM framework, namely ecological science and socially defined human wants, and the need to balance or weigh the constraints or capacities of the first against the demands of the latter. Additionally, the latter two definitions given above emphasize the dynamic or process-oriented character of ecosystems and society.

"The use of an ecological approach to achieve the multiple use management of national forests and grasslands by blending the needs of people and environmental values in such a way that national forests and grasslands represent diverse, healthy, productive, and sustainable ecosystems" (Comanor 1993).

"Ecosystem Management is the skillful use of ecological, economic, social, and managerial principles in managing ecosystems to produce, restore, or sustain ecosystem integrity and desired conditions, uses, products, values and services over the long-term (U.S. Department of Agriculture, Forest Service 1993c).

While a uniform definition of EM has yet to gain acceptance (Grumbine 1994), various definitions, including each of those listed above, do seem to have common elements; these were previously discussed in terms of the key concepts and principles of EM. In reviewing the attempts to actually define what EM means, however, one of the more striking aspects of all of the proposed definitions is their generality; although each implies a need for applied ecological science, none specifically addresses applicable ecological concepts, nor do they concretely suggest ways in which these concepts may be applied toward planning or decision making.

To understand these specifics of EM, a more detailed review of the pertinent literature proves necessary. And while the concept of naturalness seems studiously avoided in attempts to define EM and list its key concepts and principles, naturalness appears to emerge as a critical underlying concept nonetheless.

"Naturalness" emerges as an issue in Ecosystem Management at several different levels, although not often addressed in a direct attempt to define a "natural" condition or process. Within the EM framework, naturalness is usually defined as either a "pre-industrial" state (Hayes, Riskind, and Pace 1987; Kilgore 1987) or a state, within North America, preceding the effects of European American colonization (Johnson et al. 1994; Swanson et al. 1994).

Perhaps more importantly though, the term "natural" is often used in conjunction with other concepts or terms central to defining EM goals and the development and implementation of EM management strategies. The most prominent of these are concepts of: natural systems and areas, natural processes, a natural range of variability, natural disturbances, natural biodiversity, natural or native species, and defining a relationship between humans and nature.

A review of these concepts, in light of both their broader applications for ecological science and their application within the EM framework, serves to underscore the importance of naturalness as a concept both in Ecosystem Management, and in resource management in general. Additionally, such a review also points to some of the difficulties which arise in incorporating the concept of naturalness within a management framework, particularly in regard to the conflicting perceptions and assumptions which underlie management itself.

As with any other ecological application of the term "natural," identifying an ecosystem or area as "natural," in some absolute sense, poses many practical difficulties (Shrader-Frechette and McCoy 1995; Wagner and Kay 1993). Nevertheless, there remains an acknowledged need for some "reference state" in ecological studies, be it of a concrete (Barrett, Van Dyne, and Odum 1976) or more theoretical (Pickett and McDonnell 1993) character, and ecosystems or areas described as "natural" are generally so employed.

Perhaps most simply, Odum (1989) defined natural ecosystems as those "self-organized, without energy input or control from humans." Such a definition seems at first glance to be quite simple; however, it is only as simple as is our conception of ecosystems and ecological processes themselves. Given a simple conceptualization of natural systems as always being in or approaching a state

of climax, one might find the description of natural systems relatively straightforward--yet such a simple view has been replaced by a far more dynamic view of ecosystem processes, particularly in regard to succession (Botkin 1990). And as simpler theories of succession have evolved to encompass a more dynamic view, it has become problematic to describe a natural ecosystem, since their composition is itself seen as dynamic (Sprugel 1991).

Given our current dynamic view of ecosystems, critics have pointed to the idea of "natural ecosystems" as consequently being only vaguely definable, or definable only as subjective "value judgments" (Shrader-Frechette and McCoy 1995). Such observations have led some critics to dismiss the whole idea of "natural" ecosystems as a "myth" based in an antiquated "balance of nature" view of ecology (Kay 1994).

In this view, all landscapes are "cultural" rather than "natural," and areas or ecosystems which might be used as baseline or reference areas are best seen as "relics of earlier types of land use," rather than relics of natural systems (Williams 1993). Ecology in general, and ecosystem management and restoration in particular, are thus seen as being left without any real ecological reference points. Both research and management actions must therefore be based solely on evaluation framed by social and economic values, such as attainment of "efficiency" (Sullivan 1981) or creation of "useful," "desirable," or "valuable ecosystem states" (Burgess and Sharp 1981).

However, these arguments are themselves far from new, and the link between the use of natural systems as reference points and a belief in "static" nature is tenuous at best. Over twenty years ago, Egerton (1973) pointed out that the "balance of nature" concept had largely lost favor by the 1930s, today existing at most in a heavily qualified context of limits. Although those rejecting the reality of natural ecosystems seem focused on a static model of such systems, it is indeed difficult to find adherence to a static model among those who propose to study natural ecosystems.

Contrary to the argument that a concept of natural systems is based in a static view of ecosystems, it seems widely recognized by ecologists that at most there exist short-term equilibrium states (Sprugel 1991) or states of dynamic balance characterized by the variability entailed in concepts of resilience, persistence, and resistance (Pimm 1991). Certainly the majority of ecologists who view natural systems and areas as both "real" and as the logical and necessary reference point for ecological study explicitly hold to a dynamic, systems-based paradigm of ecology (Berger 1991; Franklin, Frenzen, and Swanson 1988; Kelly and Harwell 1990; Maser 1990; Norton 1987; Odum 1989; Smith and Theberge 1987).

A more serious objection to the characterization of systems or areas as "natural" rests in the "end of nature" philosophy (McKibben 1989), which asserts that "real" nature no longer exists due to the ubiquitousness of human influences. As Cain (1966) surmised, it can be observed that: "No natural area has more than a degree of naturalness." This argument is most problematic in light of the prevalence of "subtle human effects" (McDonnell and Pickett 1993) which are difficult to identify. If human influences upon ecosystems are widespread and historic to the degree that they cannot be clearly differentiated, then, some conclude, it is more logical to simply assume them as part of the "natural' character of those systems (Williams 1993).

On the other hand, however, it has been argued with equal force that no area has more than a degree of human influence (Sukopp 1976; Van der Maarel 1976), and that the impact of many subtle and historical human influences should not be treated as equivalent in scale to more recent planetary-level effects such as human-induced climate change (Boyden 1993; Peterman 1980). From this contrasting perspective, human influences are thus recognized as inherent in human ecosystems, but are still distinguished from the greater ecosystems upon which they are imposed.

First, there is the perspective that, given that human cultural practices have had far-reaching impacts on ecological systems on most parts of the earth for at least several thousand years, we should simply regard all human practices as inherent parts of those systems.

Second, there is the alternative view, that although it is often difficult to identify all effects of human cultural practices, that it is worthwhile attempting to "disentangle" human effects from the underlying ecosystems in which they occur (Castilla 1993). Without belittling the problems inherent in specifying systems or areas which are relatively "natural" in character, it is thus recognized that, without such an effort, ecology is left without a "reference state" against which to evaluate anthropogenic influences (McDonnell and Pickett 1993).

In sum, although at one level accepted by many as a reality underlying the study of ecology (Odum 1989), it is nevertheless important to note the skepticism which arises for some at the mere mention of "natural ecosystems." However, even while maintaining that "nature" unaltered by humans is basically non-existent, critics of the "myth" of undisturbed nature often nevertheless call for preservation of "natural areas" to serve a baseline function (for example, see Botkin 1990; Wagner and Kay 1993), or point to the need to "observe nature" (Turner 1993) or "emulate" "natural landscapes" (Morrison 1987) in planning the management of "natural" resources.

In essence, even while arguing against a static or "pure" vision of naturalness, most ecological researchers seem to conclude that the concept of natural systems remains viable, if only in a "dynamic and flexible context" (Johnson and Agee 1988). Although such distinctions may in this light be at times imprecise, identification of natural systems and areas is nonetheless seen as both possible and necessary in providing ecologists with the reference state or baseline needed to project or study the effects of human management (Maser 1990; McDonnell and Pickett 1993) and to rehabilitate or restore important habitats (Fosberg 1966; Herricks and Osborne 1985).

In general, it may be argued that natural systems and areas, maintained and studied in their entirety, are the only way to understand the complexity of ecosystems; in this sense, natural areas are equated with "ecological reserves," which must be sufficient in both spatial and temporal terms to sustain the complex "suites of ecological functions" of which whole ecosystems are comprised (Maser 1990). The protection of such natural area reserves, it has been suggested, need not be to the complete exclusion of human activity (Myers 1994), and the parameters necessary for such protection are seen as not easily defined (Shrader-Frechette and McCoy 1995). Nevertheless, without their protection, it may be argued that the whole study of ecology itself becomes problematic.

For implementing an EM strategy, an understanding of "the interrelationships of nature" is considered an important focus (Iverson 1993), reflected in the basic premise that EM must seek to "lace" societal wants with the "natural capacity" of ecosystems (Bormann et al. 1994). Likewise, while acknowledging the primacy of serving social wants, the maintenance of large-scale processes characteristic of "natural ecosystems" is considered a fundamental principle underlying ecological analysis and decision making (Kaufmann et al. 1994). To this end, natural systems serve not only an experimental role as a baseline for management planning, but also a functional role in actual management of non-natural or semi-natural areas as well.

Specific to the concepts of EM, "natural areas" have been cited as serving a number of functions: (1) providing reference areas, especially for assessing uninterrupted long-term ecological processes, in view of the intentional "light hand" of management; (2) providing for coarse-scale analysis of ecological pattern and process, assuming large acreages, such as designated wilderness, are set aside; (3) serving as species refugia, potential sources of biodiversity preservation, balanced against the more homogeneous character of many managed areas; (4) serving as sinks and sources of energy, nutrients, and genetic stocks for surrounding managed lands (U.S. Department of Agriculture, Forest Service 1993b).

Thus, while EM can be seen primarily as a framework for the management, and thus manipulation of ecological processes (Hammond 1993), it also indicates a need for protection of some existing "natural" areas or systems from such manipulation as well.

While definition of "natural" systems or areas must be viewed in a flexible manner (Johnson and Agee 1988), their conceptual value in defining sustainable management practices, and thus determining the goals of EM, seems somewhat inescapable. The key conceptual needs to understand a "natural range of variability," and "natural disturbance regimes" for successful management of any given ecosystem are predicated by a need to identify--and maintain--natural systems from which these can be determined (Botkin 1990; Swanson et al. 1994). Further, maintenance of nominally "natural" systems or areas can be seen as not only a basic necessity for planning and evaluating EM strategies, but also as a necessary part of long-term management, given the complex, dynamic character (and inherent unpredictability) of many ecological processes (Scientific Integration Team 1994).

In its most general terms, the concept of the "ecosystem" is itself based in an emphasis on process (Golley 1993; McIntosh 1985). The role of process is in fact so central to the ecosystem paradigm that ecosystems themselves may be characterized by the system properties which emerge from the sum of their intrinsic processes (Schindler 1988). In brief, this viewpoint holds that if we accept the premise that the component parts of ecosystems are dynamic in composition, then the only way to identify a particular ecosystem as an entity which persists over time is in terms of these process-derived properties (Jantsch 1981; Maturana 1981).

In this way, the "identity" of an ecosystem is "autopoietic" (Jantsch 1981), that is, self-defined by virtue of its intrinsic processes. If we accept the idea that the actual composition of ecosystems are constantly changing, then this emphasis on process becomes the only way in which we can meaningfully identify or discuss ecosystems as persisting through time (Jantsch 1980).

In a more applied sense, E.P. Odum (1983, 1989) has suggested that, given their generally complex character, an understanding of natural ecosystems must be based on models, and that such models, although they are necessarily simplifications of reality, may be sufficient if they are able to identify the greater "key factors" which represent the system's overall compositional processes.

Odum has further proposed that the such ecosystem models are best characterized using the "black box" model--an acknowledgment that, although we cannot know all the details of the system's parts, we can nonetheless observe its inputs and outputs (Odum, E.P. 1983). Given such an approach, ecosystems may then be understood or modeled in terms of five general classes of components: properties, forces, flow pathways, interactions, and feedback loops (Odum 1989)--natural processes and the "ecosystem fluxes" (Waring 1988) which they engender.

Perhaps the simplest approach to identifying categories or kinds of natural processes is a focus on the various mechanisms of succession (Connell and Slatyer 1977), with succession conceptualized as the sum of changes in species structures and community processes stimulated by any autogenic or allogenic forces (Odum, E.P. 1983). Once conceived in a simple linear or deterministic fashion, current theories of succession now encompass a complex variety of theoretical mechanisms and "multiple pathways," but the basic concept remains central to an understanding of how ecosystems function (Agee 1993; Christensen 1988).

successional processes constitute only one class of ecosystem processes, and a wider consideration of all ecosystem processes is also of importance, especially in respect to addressing resource management concerns (Kelly and Harwell 1990).

Simmons (1982) identified "basic ecosystem processes" as including: energy flows, mineral and nutrient pathways, population dynamics, "ecosystem changes in time," and biological productivity. Similarly, Murray (1996) addressed the issue by suggesting four primary classes or types of "natural processes": trophism, gene flow, migration, and disturbance.

Using a slightly different approach, Miles (1979) grouped the processes driving ecosystem fluctuations into "physiographic," "climatic," and "biotic" processes, while Spellerberg (1991) proposed the categories of: "slow processes" (such as climate change, succession and vertebrate population cycles), "rare events" (such as fire or volcanic activity), "subtle processes" (acid rain, biogeochemistry processes), and "complex phenomena" (intricate biotic relationships).

A more comprehensive and specific listing was also offered by Spellerberg (1991), who lists biological processes amenable to monitoring activity as including: productivity, litter accumulation, decomposition, consumption, carbon and nitrogen fixation, respiration, colonization, succession, and bioaccumulation.

Obviously, many such classification schemes are possible, and none is apt to be comprehensive for all systems in all potential situations. A key point, however, is that while successional processes merit special attention, these other processes--including a variety occurring at the levels of organism, species, community, or landscape--are of equal importance in understanding the functioning of ecosystems. Another key point is that, in applying such ecological knowledge to management, it is necessary to attempt to differentiate between inherent ecological processes which are "autogenic" versus "allogenic" (Odum 1989), or "natural" versus "management-imposed" (Simmons 1982). Otherwise, ecologists would lack any "reference state" against which to evaluate change over time imposed by human activity.

Ecosystem Management, like ecology in general, clearly has a focus on ecosystem process as a way of understanding ecosystems, and thus as a way of planning for their management (Harvey et al. 1994; Johnson et al. 1994). Reviewing various proposed guidelines or principles for EM, the emphasis on process is rather striking. For example, Kaufmann et al. (1994), in proposing a six-part list of "guiding principles" for EM, specifically address the importance of ecosystem processes in one principle, and refer to specific kinds of processes in two others.

An alternative proposal for an EM framework similarly emphasizes the importance of process in two of its four principle guidelines (Scientific Integration Team 1994), and the study of processes that have maintained particular ecosystems through time has been cited as "the best foundation" for determining the potential future conditions of managed lands (Thomas 1995).

This process orientation is also reflected in slightly less direct terms throughout the EM literature, such as in discussions focusing on the importance of natural ecosystem "function" (Kaufmann et al. 1994; U.S. Department of Agriculture, Forest Service 1994), ecosystem "change" (Bormann et al. 1994), and the dynamic and evolutionary character of ecosystems (Jensen and Everett 1994; Scientific Integration Team 1994).

The incorporation of hierarchy theory, attention to spatio-temporal scale, and attention to landscape pattern, are also aspects of EM which can be seen as ways of approaching an understanding of ecosystem function, or process through time (Bourgeron and Jensen 1994). Indeed, the dynamic character of ecosystem processes is seen as one of the greatest difficulties inherent in development of a monitoring strategy for EM (Bourgeron and Jensen 1994), and is the motivational factor behind calls for an "adaptive management" approach (Pell 1994; Walters 1986).

Perhaps most importantly, much of the EM literature explicitly recognizes that there is a need to differentiate processes which are "natural" from those variously described as "cultural" (Scientific Integration Team 1994), "managed" (Kaufmann et al. 1994; Oliver, Irwin, and Knapp 1994), "human-induced" (Johnson et al. 1994), or "artificial" (Franklin, Frenzen, and Swanson 1988), as a basis for the monitoring and on-going assessment steps which are seen as integral to the EM approach (Jensen and Everett 1994; Kaufmann et al. 1994; Scientific Integration Team 1994).

In its simplest form, the actual implementation or practice of EM can itself be described as the effort to regulate ecosystem structure and function (Agee 1994), and in this regard the manipulation or mimicking of "natural processes" are often cited as a key both to restoring overall ecosystem function following disturbance (Franklin, Frenzen, and Swanson 1988; Hayes, Riskind, and Pace 1987) and as a way of effectively manipulating systems to achieve specified management objectives (Bradshaw and Chadwick 1980; Jensen, Bourgeron, Everett, et al. 1994; Luken 1990) or more generally create "desired future conditions" (Thomas 1995).

As discussed above, natural processes cited in the EM literature are often those related to the overall process of succession, and management methods for establishing desired ecosystem states have likewise often been described as creation of "artificial succession" (Sullivan 1981) or "directed succession" (Luken 1990).

Luken (1990) suggests that such management methods for directing successional processes may be classified into broad categories of "designed disturbance," "controlled colonization," and "controlled species performance," including such specific techniques as harvesting, burning, soil scraping, tillage, seeding, planting, fertilizing, herbicide application, mowing, and irrigation.

Specific to the EM literature, potential management actions cited within this context of altering successional patterns include practices such grazing, mining, timber harvest, roading, pest management, suppressing or manipulating natural fire regimes, and various silvicultural activities (Oliver, Irwin, and Knapp 1994).

To some extent, the EM approach also recognizes that such management actions cannot precisely mimic all the effects of natural processes, for example, in substituting timber harvest for natural fire (Foss 1996; Greenlee 1995; McConnell et al. 1994), but even in these cases, an EM approach places importance on comparing management actions against natural processes as a way of assessing their ultimate impact upon overall ecosystem function and resiliency (Kaufmann et al. 1994).

Likewise, other natural processes are also seen as being of particular interest within an EM framework, even though their manipulation or mimicking may not be directly related to management prescriptions. These include biotic, chemical, and mechanical processes which comprise the hydrologic (U.S. Department of Agriculture, Forest Service 1993a; U.S. Department of Agriculture, Forest Service 1994b), edaphic (Harvey et al. 1994; Powers 1989), and evolutionary or species/population viability (Pell 1994; U.S. Department of Agriculture, Forest Service 1993b) aspects of ecosystems. The EM approach explicitly recognizes that an understanding of these and other natural ecosystem processes are necessary for sound planning and management (Agee 1993; Bormann et al. 1994; Kaufmann et al. 1994; Scientific Integration Team 1994).

The concept of an ecological natural range of variability is based in concepts of ecosystem dynamics, dynamics of both composition and function. Ecosystems are seen as varying in both spatial and temporal terms due to a combination of intrinsic processes and extrinsic environmental forces, resulting in a "natural variability" which makes it difficult to define or predict system states in a precise fashion (National Research Council 1986). In other words, natural systems by definition must be described in ways that account for variability, yet that same variability makes their description rather elusive.

Factors inducing variability include those related to overall climate, weather events, disturbance, seasonal climate changes, and cycles in organism life histories (National Research Council 1986). A major difficulty in describing natural variability is that variation can occur on a wide variety of temporal scales; in this light, one suggested way to classify such factors is in categories of: diurnal, seasonal, long-term cycles, successional, and random fluctuations (Usher 1973).

Within the ecological literature, the range of natural variability may be equated with other terms such as "historical range," "prehistoric condition," "pristine state," and so on, denoting a range of possible ecosystem conditions or states and the disturbance regimes associated with them (Swanson et al. 1994). Some attempt to take a strict view of the natural range, as what "would exist if man never trod the earth" (Inhaber 1976; Lugo 1994), while others take a more moderate stance, recognizing the lack of such truly pristine sites.

For the purposes of most North American researchers, the natural range is often equated with conditions existing prior to settlement by Europeans (Noss and Harris 1986; Swanson et al. 1994), or more generally with conditions prior to the introduction of certain cultural practices or technologies, such as industrialism or agriculture, which have radically altered ecosystem composition and function (Botkin 1990; Hayes, Riskind, and Pace 1987; Van der Maarel 1976).

This concept of a natural range of variation finds application in at least three distinct aspects of ecological science: First, variation is itself seen as a subject for ecological research. Since ecosystems are seen as dynamic, their description and understanding requires specific attention to these dynamic qualities (Ulanowicz 1986). In this perspective, change is itself seen as the proper subject of ecology, in that whole complex systems can only be defined in terms of processes (Allen and Hoekstra 1992) and that landscape patterns must be described in terms of the processes to which their development was linked (Wiens 1995).

Second, natural variation is seen as a variable which must be accounted for in the design of any ecological research or monitoring research (Goldsmith 1991; National Research Council 1986). Natural variation is seen to some extent as an unsolvable variable, and hence must be regarded as one which will always serve to limit the precision of predictions made of ecosystem behavior (National Research Council 1986). To the extent that some levels of natural variation may be described, it must be appropriately addressed in study designs, recognizing that long-term cycles or patterns may be obscured by short-term fluctuations, or conversely, that short-term fluctuations may be masked by the apparent constancy of more long-term patterns (Goldsmith 1991). Natural variation also requires recognition that the selection of any area or system state for "baseline" or reference purposes is to some degree always an arbitrarily selected point within a broader range (Hellawell 1991), rather than a truly representative one.

Third, natural variation is seen as an important indicator of the "assimilation capacity" (Izrael 1992) of a given ecosystem, that is, the range of natural variability is seen as indicating the limits of "ecological reserves" (Trojan 1984), which define the resiliency of ecosystem functions. In this way, study of the full range of natural variation is seen as a key to predicting the effects of anthropogenic impacts or assessing the long-term impacts of proposed management (Allen and Hoekstra 1992).

Specifically, a description of natural variability is often analyzed simply, in terms of the areal extent of specific vegetation types (Puchbaurer 1993). Alternatively, the natural range of variability may be more complexly interpreted through patterns of both ecosystem states and patterns of disturbance regimes (Swanson et al. 1994), through system trophic patterns (Reiners 1988), or through system flows of material and energy (Ulanowicz 1986). Characterization of the natural range of variability through description of processes or system flows proves far more difficult than doing so simply through description of community or landscape composition. However, it has been pointed out that this more difficult approach is necessary, since we are otherwise attempting to describe dynamics in static terms (Ulanowicz 1986).

The complexity of dynamics expressed by the natural range of variability is such that variability may be considered as much an "obstacle" to understanding ecosystems as it is an approach to that understanding (National Research Council 1986). This inherent difficulty in studying or describing ecological variability is in large part a result of: the broad temporal and spatial scales at which such variability may occur (National Research Council 1986; Usher 1973); the fact that processes inducing system variability may likewise interact across these broad scales (Wiens 1995); and the fact that many ecological fluctuations appear stochastic, or random, in character (O'Neill et al. 1986; Usher 1973). In light of these difficulties, two major considerations arise in applying the concept of natural variation: pattern and scale, and randomness.

Levin (1992) suggests that the key to understanding and predicting the behavior of ecosystems lies in identifying patterns and the mechanisms which underlie them. Attention to pattern becomes necessary in ecology due to the great complexity of ecosystem processes and the need to "aggregate" that complexity into data feasible for observation (Ulanowicz 1986). In this regard, it has been suggested that ecological patterns can be thought of as "persistent process clusters" (Allen and Hoekstra 1992), or patterns which

persist through fluctuation sufficiently to be identifiable. However, what is discernible as a pattern is largely determined by the scale at which a system is studied (Gaston 1993; O'Neill et al. 1986). For example, a trend in population decline observed locally may not hold true for that species at a higher level of metapopulation, or, conversely, a substantial decline at the metapopulation level may not be apparent in a study aimed at a much smaller subpopulation.

In other words, the identification of ecological pattern is made problematic by the equally problematic issue of selecting and defining scale. The selection of scale has been described as "the central problem in ecology" (Levin 1992; Wiens 1995), and it raises specific difficulties which pertain to ecology in general and the identification of natural variation in particular.

Levin (1992) points out that patterns must be understood either as emerging from the collective impact of events occurring on lower levels of scale, or as being imposed by events occurring on higher levels than those at which the patterns themselves occur. It has also been suggested that the mechanisms determining ecological pattern may in fact be interrelated across levels of scale (Wiens 1995). In this sense, the selection of a certain level of scale for study not only somewhat predetermines which patterns are available for observation, but also potentially excludes relevant information. Additionally, it must be recognized that no particular level of scale for study or description of ecosystems is inherently "proper," and the level of scale selected for any specific study must be seen as imposed by the researcher (Levin 1992; O'Neill et al. 1986).

These problems raise two additional concerns related to selection of scale. First, the problem of "autocorrelation," meaning the partial predetermination of a study's results by its selection of scale (Gaston 1993; Usher 1991), is introduced. Selection of a specific scale for studying an ecosystem will, to an extent, predetermine which patterns may be or not be observed. This can be especially troubling in identifying a range of natural variation, in that the temporal scale used in observation may in itself define that range (Gaston 1993).

Second, the difficulties inherent in selecting scale and identifying patterns of natural variation can be tied to the problem of "pseudodesign" (National Research Council 1986), or "empty correlation" (Pomeroy, Hargrove, and Alberts 1988), referring to the impossibility in ecological studies of establishing a "baseline" or control which represents true replication in an experimental sense. An important aspect of identifying a system's natural range of variation is the use of that range to define system limits or capacities (Izrael 1992; Trojan 1984), but the use of an in-part arbitrarily defined range introduces yet another level of uncertainty into an already approximated study element.

In addition to these issues of pattern and scale, there is also the issue of stochasticity in ecosystem processes. Although many variations in ecological systems may be described as cyclical or successional, others appear, at any observable temporal or spatial scale, to be random, and this kind of random variability may be seen as affecting all biological observations (O'Neill et al. 1986; Usher 1973). While this element of stochasticity may in part be accounted for through careful study design (O'Neill et al. 1986), this is not the case when the subject of study is itself a broadly defined natural variability. In particular, this element of randomness becomes a point of contention in that most ecological sampling by design assumes randomness for statistical purposes (McIntosh 1985), yet variability is conceived as having elements of pattern as well was randomness (Usher 1973). The specification and study of natural variation, then, becomes a matter of studying conjoined elements of randomness and pattern, a situation which results in a formidable task for analysis.

A formally proposed definition for the natural range of variability within the EM framework describes it as: "the composition, structure, and dynamics of ecosystems before the influence of European settlers" (Swanson et al. 1994). The relevance of this concept for EM is derived not only from fundamentals of ecosystem ecology (Odum 1983; Schindler 1988), but also from application of

the principles of landscape ecology and hierarchy theory, and the resulting conceptual approach of understanding ecosystems through an understanding of patterns of composition and process, across various spatio-temporal scales (Bourgeron and Jensen 1994; Jensen and Everett 1994).

The concept of a "natural range of variability" is recognized as an important issue within the EM framework (Grumbine 1994; Jensen and Everett 1994; Kaufmann et al. 1994), and identifying patterns of natural variability is considered a key to understanding ecosystems sufficiently to manage them on a sustainable basis (Bourgeron and Jensen 1994). It has been suggested that "natural variability" may be recognized in this context as basically synonymous with similar terms describing a range of conditions, such as "historical," "pristine," "prehistoric," and "primeval" (Swanson et al. 1994), and that the concept may, in effect, be used to define "ecosystem health," equating health with maintenance of a natural range of states (U.S. Department of Agriculture, Forest Service 1993a).

Specific to implementing an EM strategy for resource management, it has been suggested that the natural range of variation may be used as goal in itself, in that management actions should aim to maintain a representative mix of ecosystem types "across their natural range of variation" (Grumbine 1994). This is best viewed, not in a strict sense of returning whole ecosystems to some prescribed state, but rather as an overall goal of balancing other management objectives against maintaining or restoring ecosystems to their defined natural range (Swanson et al. 1994).

Less directly, it has also been suggested that the natural range of variability should serve as a "template" for developing management proposals (Oliver, Knapp, and Everett 1994), or that EM should strive to prescribe management while maintaining ecosystems within that range of conditions (Iverson 1993; Kaufmann et al. 1994). Conversely, it has also been suggested that managers need not always manage within the range, but should use it as a means of predicting possible impacts and consequences of planned actions (Leavell 1993; Shlisky 1994), or that management should be guided by maintaining ecosystems capable of

returning to within their natural range (Kaufmann et al. 1994). In either case, natural variability can be seen as providing a basis for management prescriptions and as a reference point for monitoring of system states and the impacts of management (Swanson et al. 1994).

EM proponents also have recognized a number of practical limitations that mark efforts at defining the range of natural variability for a particular ecosystem; for example, it is often difficult to interpret evidence of past ecosystem states, and it is often difficult to precisely attribute cause and effect to contemporary changes, and thus to distinguish between "natural" and human-caused states (Swanson et al. 1994). Basic problems also arise in defining temporal and spatial scales for study, especially given the fundamentally different time scales involved in human resource management as compared to geological or climatic process which have shaped ecosystems (Schindler 1988). Recognition is also given to the problem of needing to include consideration of the inherent stochasticity of some ecological processes when discussing a natural range (Kaufmann et al. 1994; Leavell 1993). The problems associated with issues of scale selection do not seem explicitly to be tied to discussions of natural variability within the EM context, but problems related to scale and pattern are addressed elsewhere within the EM literature (Bourgeron et al. 1994; Turner et al. 1994).

Despite the difficulties encountered in defining a natural range of variation, conceptualization of a natural range seems key not only to implementation of an EM management strategy, but also to definitions of ecosystem "health" (Schaeffer, Herricks, and Kerster 1988; U.S. Department of Agriculture, Forest Service 1994) or "integrity" (Anderson 1991; Reiger 1993). Further, despite the problematic nature of precisely defining a natural range, this approach must be acknowledged as a conceptual improvement over past reliance on more static approaches to judging baseline conditions (Botkin 1990; Trojan 1984) or establishing simple indicators of change (Landres, Verner, and Thomas 1988). Although slightly differing interpretations of how the concept of a range of natural variation should be applied within the EM framework, its overall applicability and importance seem well accepted in this regard.

Directly related to the concept of a natural range of variability is the twin concept of natural disturbance, which may in part be used to define the range of variability (Comanor 1993; Swanson et al. 1994) or serve as a means of judging the ultimate impact of management actions or mitigation efforts (Inouye 1988). Natural disturbance is viewed as a causative force in the process of succession (Christensen 1988), and therefore as a critical factor in the temporal and spatial patterning of ecosystems (Denslow 1985), and in maintenance of landscape diversity (Niering 1987). Nevertheless, even those who stress the importance of understanding disturbance as an integral process of ecosystems tend to acknowledge the difficulty of defining disturbance in ecological terms (Agee 1993).

Mechanisms which can initiate succession or otherwise substantially alter ecosystems have been variously described as "disturbance," "perturbation," or "stress," and although these terms are sometimes taken as being synonymous (Kelly and Harwell 1990), more commonly important distinctions are made between them.

Perturbation is sometimes used as a more general term than either disturbance or stress (Bender, Case, and Gilpin 1984), but a somewhat standard means of distinguishing disturbance from perturbation has been as follows: perturbation is taken to mean "any change in a parameter that defines a system," while disturbance is defined as "any relatively discrete event in time that disrupts ecosystem, community, or population structure" (White and Pickett 1985). In this sense, perturbations are narrowly defined as exogenous or allogenic, and the term is useful only in experimental cases when a system's structures and functions are well understood and the character of the perturbation itself is under experimental control (White and Pickett 1985).

Stress may be distinguished from disturbance in that disturbance is seen as entailing discrete events, resulting in loss of ecosystem components or structure, such as loss of biomass, while stress indicates conditions which impede ecosystem productivity or function, such as nutrient deficit (Schindler 1988). This kind of distinction has also been described in terms of "press" versus "pulse" events, in which a press event consists of some steady, sustained disruption, such as a shift in species competition patterns, in contrast to a pulse event, a relatively short-duration disruption which quickly alters species density or composition (Bender, Case, and Gilpin 1984).

In this vein, stress can be equated with "an energy drain" (Franz 1981) or "unfavorable deflections" (Barrett 1981) affecting ecosystem processes rather than more fundamentally disrupting structure. In this sense, stress can be seen as more or less the opposite of "subsidy" (Barrett 1981).

A further distinction can be made by dividing stress into two types: "eustress" and "distress" (Rapport, Regier, and Thorpe 1981). Eustress is seen as the effect of "stressors (which) may challenge the system in such a way as to evoke an adaptive response," whereas distress is seen as "a process of system breakdown entailing irreversible transformations" (Rapport, Regier, and Thorpe 1981). Also related to the concept of stress is that of "exploitation." While stress may be regarded as a kind of disruptive system input, exploitation is seen the "development of an external loop" (Margalef 1981), a particular kind of allogenic stress, such as harvesting (Usher 1973), which essentially creates a pathway that removes energy from one system or subsystem to another.

Stress, then, is related to disturbance in that both affect change in ecosystems, but stress is generally more subtle in its occurrence, interfering with ecosystem processes and straining physiological capacities (Auerbach 1981) rather than an event which directly changes the physical environment (White and Pickett 1985). Disturbance is definitively an event which instigates concrete change in ecosystem structure, and therefore initiates change in ecosystem states and dynamics as well (Acker 1990; Pickett and White 1985). While stress and disturbance are thus differentiated, the impacts of each on systems may be taken as somewhat equivalent, and much of the following discussion relating to disturbance is relevant for stress factors as well.

Disturbance can be seen as either a force internally produced, from within a closed ecosystem, or as one imposed upon an open system from without. These distinctions may be made in terms such as "endogenous" versus "exogenous" (White and Pickett 1985), "autogenic" versus "allogenic" (Odum 1989), "natural" versus "anthropogenic" (Goudie 1986), "natural" versus "human-induced" (Kelly and Harwell 1990), "natural" versus "artificial" (Sullivan 1981), or "natural" versus "restored" (Pacific Estuarine Research Laboratory 1990).

Despite the different inferences of these possible choices in terminology, all acknowledge a practical need to somehow define disturbance processes which are seen as integral to "natural ecosystems" (Agee 1993) as opposed to ones which result from either intentional of unintentional human impacts.

The kinds of events which may constitute disturbance are varied. Forms of natural disturbance noted in the literature include rain, wind, and flood events; snow and frost action; erosion (wind, water, and mass movement); volcanic activity, glaciation, and other large-scale geological processes; and animal-caused perturbations, including predation, herbivory, and mechanical processes such as trampling (Vogl 1980).

Miles (1979) suggested that such phenomena initiating change in ecosystems can generally be grouped into processes which are either physiographic (for example, erosion, siltation, slope failure), climatic (storms, frost, fire) or biotic (pathogens, herbivory, predation) in character, while Niering (1987) simply divided disturbance events into categories of natural (fire, disease, animals, drought, wind, frost) and man-induced (fire, disease, animals, pollution, past land use). Luken (1990) makes a further distinction of "designed disturbance," and lists activities which include most agricultural or resource management actions, such as crop harvest, mowing, tillage, irrigation, topsoil removal, and so forth.

In addition to identifying specific events which may constitute disturbance, it is also necessary to gain some understanding of how disturbance functions. White and Pickett (1985) suggested a number of "descriptors" as ways to evaluate disturbance events, listed in table 3.

Others have additionally suggested descriptors related to the idea of "severity," such as "recovery rate" (Runkle 1985), or "rate of patch formation" and "patch life span" (Harrison and Fahrig 1995). Another factor, similar to White and Pickett's (1985) concept of "synergism," is the concept of "temporal patterning," in which disturbances may be classified as either "phased" or "unphased," depending upon whether the disturbance regime of a particular stand or patch is partly dependent on the seral stage of adjacent areas (Abugov 1982).

Descriptor	Definitions
Distribution	Spatial pattern, relationship to topography and environmental gradients.
Frequency	Number of events over a given time period.
Interval, cycle, or turn-over time	Duration of the disturbance event.
Predictability	A function of variance in the return interval.
Area or Size	Extent of area disturbed.
Magnitude: Intensity	Force of the event.
Severity	Impact on organisms.
Synergism	Cumulative effect with other disturbances.

Much as in the definition of a natural range of variability, defining "disturbance" in general can be difficult given the "background variance" of dynamic systems (White and Pickett 1985). The problem of defining "natural" versus human-induced disturbance is in turn especially troublesome, given the problem of widespread "subtle human effects" (McDonnell and Pickett 1993) and the philosophical and practical issues attending distinguishing between human and non-human interactions (Williams 1993).

Categories of subtle human effects include: indirect effects, historical effects, biological legacies and lagged effects (McDonnell and Pickett 1993). Such subtle effects may occur over broad spatial and temporal scales, yet be of an initial magnitude such that they are not easily discerned (Russell 1993). However, such disturbances may ultimately have significant effects as they become "amplified" (Prigogine and Stengers 1984) or undergo "transmutation" (King, Johnson, and O'Neill 1991) across temporal, spatial, and hierarchical ecosystem scales (Allen and Hoekstra 1992), or as their effects cumulatively approach system threshold points (May 1977; O'Neill, Johnson, and King 1989).

One approach to the description of disturbance which recognizes the difficulty frequently met with in attempting to assign cause and effect to ecosystem phenomena divides disturbances into four categories: an obvious event with obvious effect; obvious event with subtle effect; subtle event with obvious effect; subtle event with subtle effect (Russell 1993). While such a classification scheme does not make the task of identifying a subtle event with subtle effects any easier, it does at the least keep in perspective the fact that catastrophic disturbance events, while the simplest to observe, are not the only ones of interest and importance (Agee 1993).

In addition to the problem of identifying subtle effects, other complicating factors in studying the role of ecological disturbance include the inherent background variation which occurs within dynamic systems (Sullivan 1981; White and Pickett 1985), and the occurrence of "rare events," which may occur so infrequently that they are not readily observable yet have effects which determine ecosystem structure for decades or centuries (Miles 1979). All of these make the task of identifying disturbance difficult, and they make even more difficult the nevertheless necessary task of differentiating between disturbances which are natural versus human-induced (Agee 1993; Kelly and Harwell 1990; Pickett and McDonnell 1993).

Related to each of these difficulties are also the problems inherent in selection of scale and hierarchical perspective. The initial difficulty met in this vein is that the spatial and temporal scales at which one should observe ecosystem phenomena are not specified a priori, but rather are selected by the researcher (O'Neill et al. 1986), and yet mechanisms of change generally operate on different levels of scale than the patterns they initiate, and the cause-and-effect relationships of disturbance and ecosystem change must be therefore observed across differing levels of scale rather than at one simple level (Levin 1992). Because of this, what appears to be disturbance at one scale-level of observation may appear merely as an endogenic ecosystem process at another level (Acker 1990; Sprugel 1991; Turner et al. 1993). This problem can only be resolved by studying systems at a variety of specified scales (Turner et al. 1993) and by recognition of non-nested hierarchical relationships between system components (Acker 1990; Allen and Hoekstra 1992; O'Neill, Johnson, and King 1989).

Beyond identifying and describing disturbance events and their patterns, the relationships between disturbances and ecosystems is also a central interest for ecologists. While it seems obvious that disturbance directly effects ecosystem structure, and hence function, by its alteration of the ecosystem, it is also noted that ecosystem structure may conversely impact the way disturbance itself propagates through a system and the ultimate effects it may have (Turner et al. 1989).

The way a disturbance event occurs or spreads through a given ecosystem may be seen as partly determined by factors such as the character of ecosystem structure (Turner et al. 1989), the abundance and physical/biological sensitivity of specific species to that disturbance (Abugov 1982), and the redundancy of ecosystem elements or functions otherwise disrupted (Bormann 1987).

Inherent ecosystem structures and processes which may in turn effect the response of an ecosystem following disturbance include most of those associated with models of succession, and may be categorized as elements of: composition, facilitation, tolerance, and inhibition (Connell and Slatyer 1977; Niering 1987). Examples of these might include: initial composition elements, such as the presence of relic populations, seed banks, substrate, proximity of undisturbed sites; facilitation elements of migration, mutualism, favorable environmental conditions; tolerance elements inherent in species life cycles and autecological characteristics; and inhibition elements such as competition, predation, herbivory, alleopathy, nutrient deficits (Connell and Slatyer 1977; Niering 1987).

There have also been various attempts to classify the ways in which ecosystems or populations and communities respond to disturbance. Goudie (1986) suggests that some systems may be seen as more vulnerable than others to the impact of disturbance, and the degree of vulnerability may be reflected in the degree to which a system exhibits either "inertia" or "resilience."

As a further refinement in describing these system responses to disturbance, it has been further suggested that five specific perspectives are possible: "stability," in which variables maintain a pre-disturbance equilibrium; "resilience," the process over time by which variables are displaced but return to a pre-disturbance state; "persistence," the process by which variables are maintained but for which change over time is initiated; "resistance," which is a measure of the consequences of changed states for variables; and "variability," the degree to which, or range within which, variables vary over time (Pimm 1991).

Resilience has been identified as being of particular interest for ecologists, since true "inertia" or stability is, more or less by definition, not a response to disturbance (Acker 1990; White and Pickett 1985). Resilience may be further understood as entailing other components, with special emphasis given to: elasticity, hysterisis, and damping (Westman 1991). Elasticity represents the ability of systems to respond to disturbance in an adaptive way, promoting a "metastability" over time (Goudie 1986; Naveh and Lieberman 1984; Westman 1991), while hysterisis represents the degree to which different system elements respond differently to

disturbance (Westman 1991), and damping refers to the patterns of disturbance response which occur over time, recognizing that system recovery is neither always immediate nor linear in function (Westman 1991; Westman and O'Leary 1986). Because the effects of disturbance may resonate through a system over fairly long periods of time, these perspectives are all useful in describing the effects of disturbance and arriving at some understanding or disturbance's role in observable ecosystem states.

Within the realm of EM, conservation or restoration of "natural ecosystem disturbance patterns" has been listed as a guiding principle for decision making (Kaufmann et al. 1994), and the mimicking of natural disturbance, as a means to achieve restoration of natural disturbance effects, has been described as a basic tool where ecosystem diversity and productivity are impaired (Hessburg and Everett 1994). In this regard, at a most basic level the concept of ecosystem management itself can be viewed as a combined strategy of mimicking natural processes, including disturbance processes, and avoiding or mitigating undesirable disturbances (Oliver, Knapp, and Everett 1994).

Within the EM framework, disturbance has been defined as "A short term event, whether natural or human induced, that causes a significant change from the predicted pattern in an ecological system" (Comanor 1993). Nevertheless, natural disturbance is generally distinguished from human induced or "management disturbance," as a means of differentiating between disturbance events resulting from events considered within a natural range of variability and those imposed upon ecosystems by technological management (Johnson et al. 1994; Swanson et al. 1994).

Disturbance within the EM framework is often discussed in the context of "disturbance regimes," which are in turn defined as the overall pattern of disturbance described in terms of frequency, spatial arrangement, and severity of the disturbance events (Agee 1994; Swanson et al. 1994). Although there are inherent problems in describing such regimes, given the broad temporal scales at which some disturbances may occur, a suggested approach has been to focus upon field study of disturbance processes in conjunction with paleoecological and dendrochronological methods for studying past events, thus offering at least a reasonable estimation of historical patterns of regime characteristics (Swanson et al. 1994).

An alternative EM perspective on disturbance regimes suggests that they must be described in similar terms of frequency, return time, and magnitude--but also "predictability" (Christensen 1988). In this view, predictability is seen as a measure of a regime's variability, a consequence of natural variations in environmental conditions including diurnal, seasonal, and random fluctuations, all of which may influence a disturbance event's magnitude. The variability of disturbance is in this way considered a direct causative factor in determining ecosystem variability as well. (Christensen 1988).

The distinction between natural and management disturbance can be seen as important within the EM framework, in that disturbance is of interest in three different contexts: first, in the need for an understanding of natural disturbance as a way of defining an ecosystem's range of natural variability (Swanson et al. 1994); second, in the need for identifying the ways in which past management actions have interfered with natural disturbance and thus indirectly altered ecosystem function (Agee 1994; Wiens et al. 1993); and third, in the need to identify past human-induced or management disturbance and the direct effects it may have had on ecosystem function (Harvey et al. 1994; Johnson et al. 1994; Oliver, Irwin, and Knapp 1994).

As for understanding the role of natural disturbance, natural disturbance regimes are seen as having been a key determinant of natural landscape pattern, and as a key determinant of the dynamic range of variability that characterizes a given system (Bourgeron and Jensen 1994; Jensen, Bourgeron, Everett, et al. 1994; Swanson et al. 1994). Examples of natural disturbance events

which have thus facilitated ecosystem function and are cited in the EM literature include fire, grazing and browsing, insect and disease cycles, glaciation, wind storms, floods, soil mass movement, and climate change (Johnson et al. 1994). All of these factors are seen as having had a role in the development of historic landscape pattern and ecosystem function (Johnson et al. 1994; Turner et al. 1994), although some are apparently also regarded instead as forms of ecosystem stress (Jensen, Bourgeron, Everett, et al. 1994).

Discussion of management disruption of natural disturbance patterns tends to center on the topic of fire suppression (Agee 1994; Johnson et al. 1994), but management practices seen as disrupting natural disturbance regimes also include pest suppression, flood control, withdrawal of water for irrigation, timber harvest, and grazing by livestock (Hessburg and Everett 1994).

In many cases, the disruption of natural disturbance regimes are seen to have had significant impacts, while in others it has merely resulted in a delay in disturbance frequency, with subsequent increases in severity and extent, or perhaps resulted in synergistic effects on other disturbance forces (Agee 1994). For example, the disturbance regimes of defoliator insects may be indirectly altered by suppression of natural fire regimes, or vice-versa (Hessburg and Everett 1994). Restoring specific disturbance effects is thus seen as a necessary approach for reducing both direct and indirect forms of ecosystem stress (Jensen, Bourgeron, Everett, et al. 1994) and for restoring the natural range of variability in system function (Kaufmann et al. 1994).

Aside from management actions which directly attempt to alter natural disturbance regimes, many other management practices constitute disturbance in their own right. The harvest of timber and other natural resources, grazing by domestic livestock, planned burning, introduction of non-native species, and various activities disturbing soils and hydrological processes can all be described as management-induced disturbance (Johnson et al. 1994; Luken 1990). The effects of these actions are often cumulative

over both temporal and spatial scales, and in many cases are recognized to have had substantial effects on ecosystem structure and function (Johnson et al. 1994; Oliver, Irwin, and Knapp 1994). Obviously, these kinds of management practices will continue, but a greater understanding of both natural disturbance effects and the effects of management disturbance may allow for a "prudent level" of management disturbance (U.S. Department of Agriculture, Forest Service 1993a) such that ecosystems are maintained within their natural ranges (Kaufmann et al. 1994).

Within the EM framework, the protection or restoration of natural disturbance is seen as an important tool in maintaining ecosystems, by protecting biological diversity (Johnson et al. 1994) and by maintaining natural landscape patterns (Turner et al. 1994), and overall ecosystem processes (Kaufmann et al. 1994). Also, an understanding of natural disturbance regimes is seen as key to modeling management activities so as to achieve management goals within the constraints of ecosystem capacities (Agee 1994; Jensen, Bourgeron, Everett, et al. 1994; U.S. Department of Agriculture, Forest Service 1993a).

In this regard, the role of natural disturbance has also been a focus of attention within the framework of ecological restoration work. For example, much of the early work performed by restorationists dealt with prairie ecosystems, where the reintroduction of fire (which in recent history had been suppressed by human intervention) was considered the "essential tool" of restoration efforts (Kline and Howell 1987). Likewise, restoration efforts in forested areas have centered on a restoration of "natural fire process" in response to the unintended ecological effects of past human land management, such as insect and disease epidemics or episodes of severe wildfire (Arno et al. 1995).

However, a prevalent theme in the EM literature is that allowance for, or the restoration of, natural disturbance is in many instances incompatible with other management goals (Agee 1994; Christensen 1988; Jensen, Bourgeron, Everett, et al. 1994; Oliver, Knapp, and Everett 1994). Paradoxically, the maintenance of natural disturbance may be variously viewed as both a specific goal or endpoint for ecosystem management, and as an impediment to other ecosystem management goals (Agee 1993).

In fact, while analysis of natural disturbance regimes may be considered a key to defining an ecosystem's natural range of variability (Swanson et al. 1994), it has conversely been argued that, because many managed systems are now outside their natural range, natural disturbance should be considered "inefficient" for restoring or maintaining ecosystem integrity (Everett 1995). In this vein, the EM literature more frequently calls for "mimicking" natural disturbance (Jensen, Bourgeron, Everett, et al. 1994; Oliver, Knapp, and Everett 1994), manipulating natural disturbance processes as a "tool" (Agee 1994) or restoring "natural disturbance effects" (Hessburg and Everett 1994) rather than actually placing emphasis on reinstating natural disturbance regimes themselves.

Nevertheless, it is also true that our knowledge of ecological systems is inherently imperfect (Pomeroy, Hargrove, and Alberts 1988), and that, for example, while a particular silvicultural treatment might bring a forest back within its natural range of patch or stand composition, it may yet fail to truly duplicate all the effects, particularly soil conditions, which would result from a lightning-caused fire (Greenlee 1995), thus altering long-term site characteristics and function (Harvey et al. 1994).

Likewise, given the inevitable unknown factors in ecosystem function, it will often prove difficult to determine whether management actions in response to disturbance are actually necessary for the purpose of maintaining ecosystems (Foss 1996), and although activities of management may in some ways effectively emulate natural disturbance, they may also have other undesirable effects on the inherent variability of disturbance regimes (Christensen 1988). These uncertainties may prove additionally difficult to account for within a "target-driven EM" approach, which still gives an overriding attention to meeting goals of resource productivity within an EM perspective (Artley 1994).

While the difficulty in attempting to maintain ecosystem function by substituting management actions for natural disturbance processes is not seen as insurmountable by EM advocates, critics fairly point out that much if not all past management actions now recognized as undesirable forms of management-induced disturbance or stress were, at the time of their implementation, considered to be favorable and effective actions for managing ecosystems (Donnelly 1995). In this regard, EM will necessarily have to strive to both mimic natural disturbance in the course of management activity and maintain natural disturbance processes themselves where feasible (Christensen 1988; Kaufmann et al. 1994; Turner et al. 1994).

Ecological diversity has been cited as a central concern for ecology in general (Wilson 1988), for ecological restoration work (Jordan, Peters, and Allen 1988), and for EM (Glenn and Collins 1994) in particular as well. Fosberg (1966), in a discussion of the goals of ecological restoration, suggested that the "pre-eminent guiding principle" in restoration work should be "re-creating the full diversity, both of single communities and of landscapes" which occurred in "natural habitats." Fosberg wrote specifically of the need to protect and restore "natural diversity" in ecosystems, but that phrase is now often superseded by the use of "species diversity," "biotic diversity," "ecological diversity," "biological diversity" and "biodiversity" instead.

The issue of biodiversity has recently been the subject of numerous books and articles, but like many such technical terms which have entered the mainstream of public discourse, biodiversity is often given a variety of meanings. In this regard, it was suggested that "the term species diversity has been defined in such disparate ways that it now conveys no information . . . species diversity has become a nonconcept" (Hurlbert 1971). Since these comments were published, a great deal of effort had been made at clarifying the concept of diversity and incorporating it into ecological theory; however, somewhat varying interpretations still exist. As a more recent critic of the diversity concept put it: "We are all concerned with biodiversity, but with what biodiversity?" (Bradshaw 1994).

The simplest definition of biological diversity equates it with the number of species found in a given system, or "species richness" (Erwin 1991), an interpretation often used and fairly simple to apply within various assessment techniques. A somewhat broader approach is to focus on species but also examine richness at other taxonomic levels, such as genera, recognizing the inherently artificial and somewhat problematic methods used to classify organisms (Ray 1988; Wilson 1988).

Alternatively, researchers often discuss biological diversity in terms of two complimentary concepts: species "richness," but also species "abundance," recognizing that the mere presence of a species is in itself an insufficient way to evaluate viability. Both of these concepts may in turn be measured and evaluated using various sampling methods, indices, and analysis approaches (Magurran 1988).

The concept of diversity may be further expanded into the broader perspective of "ecological diversity" (Magurran 1988), with attention focused variously on diversity at the levels of species assemblages, or communities (Keddy 1994), or landscapes (Niering 1987), or broadened even further to consideration of diversity in terms of geographic, ecological, and genetic "patterns" (Wilcox 1990). In this expanded perspective, biological diversity can be thought of as constituting a system of "nested biodiversity," in which diversity at the genetic level affects diversity at the species level, which in turn affects diversity at increasingly higher levels of ecosystem structure and process (Kim 1993). In this broadest sense, biological diversity can be described as "the variety of life and its processes" (Keystone Center 1991).

Those who adhere to a more simplistic view of diversity, conceived as meaning species richness, tend to see "diversity" as an end in itself for ecosystem management, placing the maximization of species richness above the maintenance of "natural" systems or

Diversity interpreted in a dynamic and broadly ecological way recognizes that diversity, in terms of both richness and abundance, will tend to vary over both temporal and spatial scales in response to ecosystem disturbance regimes and species life cycles or population dynamics (Abugov 1982), whereas the narrower view of tends to focus attention on individual species, resulting in a limited spatial and temporal scale perspective (Noss and Harris 1986). For example, Wiens (1995) proposes that a description of landscape diversity should be expressed in terms of richness, evenness, dispersion, and predictability. In this way, we might think of the dynamic interpretation of diversity as constituting an ecosystem approach, as opposed to the static, organism-based approach of a species-richness interpretation.

Additionally, diversity may be conceptually linked with ecosystem function or process, instead of simply with composition or structure (Kim 1993; Ray 1988). Within this context, diversity is properly interpreted through "patterns" as well as quantity (Lovejoy 1988; Noss and Harris 1986), and it is seen that relative "rarity" or "abundance" do not directly correlate with importance for ecosystem function (Lovejoy 1988). Likewise, it has been observed that maximized diversity at a fine-scale habitat level may actually lead to less diversity at a coarser landscape scale, due to increasing coarse-scale homogeneity (Noss and Harris 1986), and that "ecological or functional uniqueness" may be of more concern than simply rarity in terms of population numbers (Orians and Kunin 1990).

While a simple species-quantity perspective of diversity favors maximum species number over the integrity of natural populations (Gotmark 1992), a dynamic and hierarchical perspective discounts the value of "artificial diversity" (Angermeier 1994) and places explicit importance on "natural" patterns of abundance and function (Anderson 1992; Noss and Harris 1986). Natural diversity is not implied to be the only desirable or valuable quality of managed ecosystems, but it is in this perspective recognized that "native" (Maser 1990) or "natural" (Anderson 1992) systems are in part composed of their dynamically diverse populations.

In sum, "biological diversity" is more broadly defined than simple "species diversity" or "richness," and encompasses the concept of diversity applied at levels varying from genes to species, from landscape patches to whole ecosystems (Angermeier and Karr 1994). Diversity is further seen to vary across spatial, temporal, and hierarchical scales which are not necessarily directly correlated, and, by focusing on ecosystem function versus simple richness or abundance, is seen to be a relative value which cannot be simply "maximized" in terms of quantity (Angermeier and Karr 1994).

Perhaps most importantly, diversity, when defined in the context of ecosystem function, becomes tied to concepts of "native" or "natural" communities (Maser 1990; Noss and Harris 1986) and the concept of "biological integrity" (Angermeier and Karr 1994), in that diversity is seen as contributing to the integrity and sustainability of naturally evolved systems, rather than as an end in itself.

Protection or management of biological diversity is often identified as a key component of an EM approach (Bormann et al. 1994; Hann, Keane, et al. 1994; Kaufmann et al. 1994; Marcot et al. 1994; McConnell et al. 1994;). Discussion may center on biological diversity as a parameter of EM in itself (Hopkins 1993), as a resource component of ecosystems (Hann, Keane, et al. 1994), or as tool for achieving greater overall ecosystem management goals (Bormann et al. 1994).

Specific to the EM framework, biological diversity has been variously identified as a means of ensuring ecosystem "adaptability" (Bormann et al. 1994; Everett, Oliver, et al. 1994), "viability" (Iverson 1993) or "sustainability" (Hann, Keane, et al. 1994; Marcot et al. 1994; McConnell et al. 1994). Such assessments can be seen to infer recognition of the role played by biological diversity in ensuring long-term maintenance of ecosystem function and process (Hann, Keane, et al. 1994).

As is true within the greater field of ecology, managers attempting to define EM and its goals often have slightly differing views of specifically what biological diversity entails. Some take a more narrow view of diversity, as meaning species richness and abundance, even proposing that study of select "indicator species" within this context could serve to monitor the impact of management actions (Hopkins 1993), while others focus more generally on habitat-type diversity (Everett, Hessburg, et al. 1994). However, the majority of authors seem to have adopted a more dynamic and hierarchical approach to diversity.

Marcot et al. (1994), for example, discuss biological diversity at the levels of species, populations, genomes, and communities, while others add the concepts of "metapopulations" (Glenn and Colllins 1994; U.S. Department of Agriculture, Forest Service 1993b) or landscapes and whole ecosystems (Hann, Jensen, et al. 1994). Also within the EM literature are mentions of specific concern with genetic diversity (McConnell et al. 1994) and unique, scarce, or declining habitat types (Marcot et al. 1994; U.S. Department of Agriculture, Forest Service 1993b). A definition offered for biological diversity within the EM framework describes the concept as encompassing both "the distribution and abundance of plant and animal communities," and "the variety of life forms and processes, including a complexity of species, communities, gene pools, and ecological functions" (Comanor 1993).

For the most part, those discussing the importance of biological diversity within the EM framework strongly emphasize that diversity is a factor underlying various aspects of ecosystem complexity and function (Bormann et al. 1994; Hann, Keane, et al. 1994; Kaufmann et al. 1994; U.S. Department of Agriculture, Forest Service 1993b). It has also been suggested that a focus on managing biodiversity expressly offers an ecosystem-based alternative to past approaches of single-species management (Marcot et al. 1994), emphasizing instead greater consideration of landscape-scale issues of structure and process (Glenn and Collins 1994). Hence, EM's emphasis on the importance of monitoring and managing for biological diversity can be seen as a part of EM's overall shift towards a management strategy more firmly based in fundamental principles of ecosystem ecology.

Perhaps the most significant aspect of EM's incorporation of biological diversity as a central concern is the importance placed upon viewing biological diversity not merely as a quantitative endpoint, but as a functional component underlying the "capacities," "adaptability," or "sustainability" of ecosystems. EM also explicitly accepts the premise that "natural communities are dynamic" (Glenn and Collins 1994) and that diversity in terms of both distribution and abundance of species at all levels of scale must be viewed as fluctuating over time.

The EM approach to diversity is thus seen as function-based; and it can also be seen as explicitly addressing the importance of natural diversity, as a benchmark against which management-induced changes may be compared. For example, it has been suggested that defining sustainability requires arriving at an understanding of the "natural capacity of ecosystems," and that these capacities must in part be evaluated through measures of biological diversity (Bormann et al. 1994). Likewise, the dynamic patterns of species distribution and abundance in "natural communities" have been cited as an important reference point for modeling the effects of management actions (Glenn and Collins 1994).

Similarly, it has also been suggested that, even though it is expressly not EM's intent to preserve all ecosystem states, a goal of EM is nonetheless to "preserve all components of natural ecosystems" (Kaufmann et al. 1994), or "the maintenance of natural ecosystem conditions and processes" (Hann, Jensen, et al. 1994). Preservation of genetic diversity, as well as structural and functional diversity, is seen as a crucial need specifically in intensively managed landscapes (Franklin 1988). In this sense, while the phrase ecosystem "management" in itself clearly implies alteration of natural systems for human resource extraction, it also implies the need for understanding and protecting the dynamics of natural biological diversity, simply as an aspect of maintaining overall ecosystem function.

According to Ralph Waldo Emerson, a weed is "a plant whose virtues have not yet been developed," whereas J.L. King offered the idea that a weed is even more simply "a plant out of place." (Parham and Healy 1976). Given these kinds of definitions, plants regarded as "weeds" often include native species, especially ones which invade or persist in agricultural lands (Taylor 1990). In the case of agriculture, this view of what constitutes a weed can even lead to the eradication of native plants with herbicides in order to favor non-native species (Dow-Elanco 1991). In this sense, "weeds" are defined mainly in relation to their effect on economics or commodity production.

Regarding species instead in terms of their role in ecosystem structure and function, ecologists have generally used the term "non-native" to describe organisms which are "out of place," and the term "native" for those considered "in place." In this ecological context, the term "native" is taken simply to mean "indigenous, originating in a certain place" (Wilson, Hibbs, and Alverson 1992). This term, however, is not always used in ways precisely defined, and its ecological meaning may be entangled in different meanings more related to political or philosophical contexts.

For example, some discount the impact of non-native plants, or even condemn the distinction between native organisms and humanly introduced ones as "perversely misanthropic" (Scherer 1994). Such stances against the very idea of a species being called "native" apparently see the concept as an attack on humanism (Jordan 1994), and categorize statements regarding native species or natural environments as, "insulting" (Bonnicksen 1995) or even suggest they constitute an extension of racist beliefs founded in Nazism (Jordan 1994; Marinelli 1995).

Those rejecting the distinction between native and non-native species may also take the position that human manipulation of species distributions and genetics merely fulfills humanity's role as ecological "manipulators" (Moggridge 1986; Scherer 1994), or "the ultimate evolutionary process" (Sargent 1974). From this perspective, all species occurring in any habitat, regardless of their origin, must be regarded as equally "natural," whether the result of evolutionary and geobiophysical processes or human artistic effort (Turner 1985, 1994).

The argument against the usefulness of categorizing species as native or non-native may also be framed in terms of "substitutability," the theoretical position that the loss of any species in an ecosystem can be fully mitigated by introducing some similar organism, which will then replace the lost species by filling its ecological role (Gunn 1991). Substitutability is certainly far from an accepted theory, however, and the complexity of ecosystems and species interactions are such that substitutability may be regarded more as a philosophical stance than a technically feasible strategy (Cairns 1989; Elliot 1994).

In ecological terms, moreover, it is recognized that ecosystems must be defined in terms of their composition, structures, and functions (Odum 1989), and that the description or study of a system requires identifying these component parts. In this regard, arguing against the very possibility of differentiating native and non-native species amounts to a refutation of ecosystems themselves. Tansley, who coined the term "ecosystem," also coined the terms "autogenic" and "allogenic" (McIntosh 1985), to indicate those organisms or processes which are part of an ecosystem or part of its greater environment. The ecological context of identifying "native" species is perhaps best viewed in this light; non-natives, by definition, represent interference with intrinsic ecosystem composition and function (Temple 1990).

The key to a useful ecological definition of "native" requires certain spatial and temporal perspectives (Wilson, Hibbs, and Alverson 1992). In, spatial terms, it falls to the domains of history, biology, and geography to define the limits of where any particular organism is "native," based on its distribution as measured against climatic, biological, and human-induced changes. Using this geographic context, it becomes possible to discuss species, ecosystems, and whole landscapes in terms of their native components (Wilson, Hibbs, and Alverson 1992). A species of plant, for instance, may broadly be considered native to a large area such as North America, but not to a particular subregion of that area. Likewise, a plant may be native within a particular life zone or ecotone without being native to proximate areas with differing environmental constraints, or may be native only to specific microclimate conditions.

Along with this spatial perspective on identifying natives, there is also a temporal element. The species-composition of any ecosystem will change over time, regardless of human intervention. Past plant communities in a given locale, as inferred from the fossil record, are often far different from those observed today, not only in regard to individual species, but also in regard to their associations (Brubaker 1988). Although species composition at a given location may have considerable natural variation over relatively long time periods (Brubaker 1988), historical records also clearly demonstrate that there are instances when humans have introduced otherwise alien species, causing varying degrees of ecological disturbance (di Castri 1989; Wilson, Hibbs, and Alverson 1992). Some shifts in species distribution may thus be seen as occurring due to intrinsic ecosystem processes, while others may be identifiably human-caused. These distinctions are not always evident, however, and may be complicated by the occurrence of small refugia populations, species migrations, successional changes, and other transient ecosystem processes or states (Wilson, Hibbs, And Alverson 1992).

Further complicating the identification of natives, there are also genetic considerations which call for caution in defining a particular plant or animal as being native to a given locale. In some cases, a species may be thought of as native to an area, but only as a specific local gene pool (Alverson 1993).

The process of identifying organisms as native to a given ecosystem is thus a somewhat imprecise and difficult matter. Nonetheless, it is an important concept for identifying ecosystem components and studying or predicting ecosystem behavior, and this matter of biologically determining what belongs and what does not has become a basic task of conservation biologists and ecological restorationists alike (Soulé 1990). More specifically, the concept of native species is an important one underlying two major concerns of ecosystem science and management: the biodiversity, and the integrity, of natural systems or communities.

Regarding biodiversity, the issue of native versus non-native finds particular importance. Some ecologists have argued that a broadly defined maintenance of ecosystem function is a sufficient goal for environmental management, whereas others have argued for preservation of individual species in acknowledgment of the known and potential values they represent as contributing to biodiversity (Soulé 1996).

Maintenance or perhaps restoration of native species in populations within a natural range of variation is seen as a key component of a biodiversity-based conservation strategy on several grounds: natives are seen as "co-evolved" (Lawlor and Smith 1976; Woodley 1993) elements of their ecosystems, and hence are seen contributing to both overall ecosystem function and the viability of other species with which they are associated through both competition and cooperation with other species (Augros and Stanciu 1987; Lawlor and Smith 1976).

Natives are also seen in this regard as an element of both landscape-level continuity and patch-level diversity; and as being in themselves more viable and resilient than non-natives which may otherwise be used to partially substitute for them ecologically (Handel, Robinson, and Beattie 1994). In contrast, the introduction to or invasion of an ecosystem by non-natives may have far-reaching effects on both ecosystem structure and function (Berger 1993), disrupting ecological habitat and inter-species relationships and reducing overall biodiversity through their displacement of native species (Monsanto 1995; Mulroy 1989; Padgett and Crow 1994).

Native species are also a concern as a matter of biodiversity at the genetic level, and it has been suggested in this regard that preservation of biodiversity should focus "beyond population persistence to gene persistence (Connor 1990). In this sense, natives may be defined at the "ecotype" level, recognizing that genetically differentiated strains may arise within populations which serve specific ecosystem functions below the species level (Smith 1994).

Regarding ecosystem integrity, native species are viewed individually as unique components of ecosystem composition, and the viability of their populations, at the level of community interactions, is considered a key to defining integrity itself (Woodley 1993), even when those interactions constitute a "redundancy" of function, to the degree that such redundancy occurs under natural conditions and within a dynamic range (Karr 1993). By definition, native species maintain the integrity of ecosystem function while non-natives alter it (Berger 1993).

Distinguishing native species from non-natives is also of importance in regard to the kinds and levels of impacts which non-natives may have. The introduction of non-natives is usually termed "invasion," although by definition those non-native species which have invaded indigenous ecosystems have done so because of human-caused disturbance (Rahman 1982; Wilson, Hibbs, and Alverson 1992). Biological invasions can occur at a number of ecological 9levels, ranging from microscopic plant or animal pathogens to plants, insects, or vertebrate animals (Vitousek 1986). In all cases, however, it is important to note that such invasions occur with impacts far beyond the species level, with effects often rippling across entire ecosystems (Ramakrishnan and Vitousek 1989).

For example, the microscopic fungus Cromartium ribicola, commonly known as "white pine blister rust," was accidentally brought into North America along with some nursery stock from Europe (Broembsen 1989). Over the course of just a few decades, this non-native pathogen spread across the entire continent, and has to date eliminated about 90% of the native western white and whitebark pine from their natural range (Campbell 1993). Extinction of these tree species seems a possible eventual outcome. In the western part of North America, the afflicted whitebark pine is now dying-out throughout its range, which may effectively lower timberline, changing a critical ecotone, and eliminating a primary food for numerous animals, ranging from squirrels and small birds to grizzly bears (Arno and Hoff 1989).

Similarly, a Canadian study found that in areas where alien species of vegetation had become established, native bird species were affected in terms of both species distribution and abundance--even though there were no introduced species of birds present (Wilson and Belcher 1989). The authors concluded that introduced species of plants had profoundly changed the ecosystems, even at higher trophic levels where no direct disturbance had occurred (Wilson and Belcher 1989). Likewise, the introduction of a single, competitive species of grass was shown to have altered the diversity and abundance of various insects, rodents, and birds, as well as other species of plants, throughout the Southwestern United States (Bock et al. 1986).

As to the ecological management of natural areas or natural communities, a number of approaches to natives and non-natives may be taken. Some, holding to the philosophy that all human-induced ecological changes should be considered "natural," maintain that invasions of non-natives are not in themselves a concern, and that they should be simply regarded as "nature's way," even in light of known changes in ecosystem structure or function which imperil the viability of indigenous species (Messmer 1995). Others recognize the distinction between native and non-natives, but suggest that in practical terms it may be necessary to simply distinguish between "exotic ecosystems" dominated by non-natives and "naturalized ecosystems," in which non-natives exist but do not appear to play a dominate role (Johnson and Carothers 1987), or that the exclusion or elimination of non-natives may be regarded as an ultimate goal, although relative levels of success may be necessarily regarded as sufficient within specific time frames or systems constraints (Harrington 1989).

Once non-natives are established in a given ecosystem, eradication is, in most cases, simply impossible (Berger 1993). Various management practices may be used to attempt restoration of the naturally occurring ecosystem structure and function, including mechanical and cultural methods, the use of chemical pesticides, or the use of biological controls; however, all of these methods may prove to have limited effectiveness and relatively high implementation costs when non-native populations attain any significant level of establishment (Kummerow 1992; Mullin 1992).

Direct references to the importance of native species appear to be infrequent within the EM literature, although Grumbine (1994) lists "maintain viable populations of all native species in situ" as one of five goals for EM. However, the issue of native versus non-native species seems most often to be expressly acknowledged in the EM perspective in that native species are regarded as components of ecosystem integrity and biodiversity (Grumbine 1994; Marcot et al. 1994; Woodley 1993). Additionally, It is noted that there are legal mandates for special concern regarding native species entailed in the National Forest Management Act (Marcot et al. 1994), the Endangered Species Act (Swanson et al. 1994), and the Clean Water Act (Suter and Barnthouse 1993), which must be addressed within the EM framework.

Aside from the role of native species in defining and maintaining ecosystem integrity and diversity, this issue is also recognized more specifically in the context of non-native species and the impacts they may have on long-term ecosystem function (Kaufmann et al. 1994; Swanson et al. 1994). Preventing the introduction of non-natives, promoting natural levels and patterns of species diversity, and both restoring extirpated natives and eradicating non-natives where feasible, are all seen as parts of an overall EM strategy for management (Kaufmann et al. 1994). Further, sustaining native species populations within a natural range of variability has been suggested as an EM objective in itself (Swanson et al. 1994).

In regard to the impacts of non-natives, the study of their consequences and development of mitigation measures is considered a primary research need (Kaufmann et al. 1994), and their introduction or invasion is regarded as a form of ecosystem degradation which threatens long-term sustainability of ecosystem function (Hessburg and Everett 1994). On the other hand, however, it is also recognized that non-natives are extremely prevalent in some ecosystems, and that while native species should be favored by management, the eradication of all non-natives is in many cases simply not practical (Johnson et al. 1994).

The issue of native and non-native species also raises a problematic concern for EM in regard to its heavy emphasis on monitoring and research as the basis for adaptive management. The complexity of managing or monitoring whole ecosystems has resulted in a tendency to rely on relatively simple indicators of ecosystem integrity or health, often focusing on landscape-scale phenomena (Bourgeron and Jensen 1994; Bourgeron et al. 1994). In response to this, it has been suggested that only select species may be effectively managed for, and that concern cannot be given to individual species (Bormann et al. 1994). Conversely, it has been acknowledged that the use of indicator species often fails to achieve its purpose due to the complexity of species interactions and ecosystem processes (Landres, Verner and Thomas 1988; Marcot et al. 1994).

In essence, the EM literature acknowledges that a species-by-species management approach has proven impractical and failed to protect ecosystem function (Bormann et al. 1994; Marcot 1994), but also acknowledges that each native species must be considered as important to ecosystem diversity and function. These conflicting perspectives have been somewhat addressed by the proposal for a twin approach of "coarse-filter" and "fine-filter" management (Bourgeron and Jensen 1994; Marcot et al. 1994; McConnell et al. 1994) and a flexible focus on endangered, threatened, sensitive and other "featured" organisms at genome, population, and community as well, species levels (Marcot et al. 1994). Nevertheless, the importance of individual native species to ecosystem function and the hazards posed by individual non-native species obviously afford a significant challenge to the whole-ecosystem and landscape-scale approach of EM.

An early effort at on-the-ground application of EM principles found that varying perspectives on the role of humans in relation to nature constituted an "early stumbling block" for the project's planning team (McConnell et al. 1994). And indeed, the development of an EM approach to managing natural resources has explicitly devoted much attention to the fact that the term "ecosystem management" implies both attention to ecosystem science and the purpose of managing for human purposes (Grumbine 1994; Kaufmann et al. 1994).

In this regard, EM has been described as entailing a two-way relationship of elements, meshing human social and economic wants with the bio-physical constraints of ecosystem capacities (Bormann et al. 1994), or as a three-way relationship between social human desires, technological and economic factors, and ecological capabilities (Jensen, Bourgeron, Hessburg, et al. 1994). These attempts to merge applied ecological science with management for human-centered concerns have been described as a "contextualism" model (Grizzle 1994) or "lacing model" (Bormann et al. 1994), in which human wants and the capacity of ecosystems to meet those wants are "laced" or analyzed to determine the limits to which they are mutually compatible.

In this regard, an overriding perspective of EM seems to be that of "humans embedded in nature" (Grumbine 1994), an acknowledgment of humans as both dependent inhabitants of their own ecosystems, and as a "keystone species" (Harvell 1990; Kay 1994) which has the ability to affect tremendous ecological change.

A significant part of the EM approach is based in close attention to sound ecological science (Grumbine 1994; Kaufmann et al. 1994), but while some see this role as primary in the goals of EM (Goldstein 1992; Grumbine 1994), the EM literature generally places a greater emphasis on the idea of EM as a strategy for managing resources to human benefit. From this perspective, EM itself is seen as a way of resolving social conflict over how to set priorities for management (Daniels et al. 1994). Plainly, a lack of agreement on how to respectively emphasize the ecological and social aspects of EM remains a substantial challenge, both in defining EM's goals (Donnelly 1995) and in prescribing management actions within the EM framework (McConnell et al. 1994).

Grumbine (1994), for example, suggests that "ecosystem management integrates scientific knowledge of ecological relationships within a complex sociopolitical and values framework toward the general goal of protecting native ecosystem integrity

over the long term." Also in this vein of thought, one U.S. Forest Service publication suggests that "Ecosystem management involves a shift in focus from sustaining production of goods and services to sustaining the viability of ecological, social, and economic systems now and into the future" (Kaufmann et al. 1994).

In contrast, other U.S. Forest Service publications suggest that the aim of EM is indeed to provide for "a sustained flow of values, goods, and services" (Hessburg, Haynes, and Everett 1994), or a "desired yield of goods, services, and states from an ecosystem" (Bormann et al. 1994). Likewise, it has been stated that meeting social "desires" is central to the definition of EM (Daniels et al. 1994; Iverson 1993), and even EM objectives such as protecting natural biological diversity may be seen as driven by the larger goal of needing to "maintain a high quality existence for humanity" (Iverson 1993).

Despite an emphasis on applying ecological science to decision making, much of the EM literature also clearly focuses on the need to "recognize political realities" (Everett, Oliver, et al. 1994), even suggesting that "market research" is a necessary part of developing EM objectives (Bormann et al. 1994). In fact, EM is generally seen not merely as an ecological approach to resource assessment, but as a management system in itself (Bormann et al. 1994; Scientific Integration Team 1994); the difficulty in defining the nature/humanity relationship in EM thus amounts to a problem of maintaining some emphasis on ecosystems and ecosystem ecology within what is actually a management system per se.

Viewing EM specifically as a management system, Slocombe (1993) distinguishes several very general steps in implementing an EM approach: defining the management unit; developing an understanding of the resources involved; planning; and creation of a management framework. In a somewhat more refined fashion, Van Riet and Cooks (1990) suggest a planning model based on the following steps: development of objectives; inventory of perceived resource values; ecological landscape analysis; evaluation, development of alternative proposals; comparative analysis; and zoning or land allocation.

Like Slocombe, however, Van Riet and Cooks fail to propose a detailed plan which includes decision making or actual management. In fact, a key assumption of many scientists promoting an ecosystems approach to resource management is an acknowledgment that EM, while providing a framework for management, cannot also serve to establish goals which are more properly set in a sociopolitical realm (Grumbine 1994). Even in efforts to frame EM as a broad process including decision making (consisting of the four iterative steps of Assessment, Decision making, Implementation, and Monitoring), the actual process of making decisions is nevertheless rather vaguely defined, deferring in large part to the separate and legally superseding realm of the "NEPA" process and its attendant regulations (Scientific Integration Team 1994).

In fact, while the EM framework considers ecological science, along with social science, and economic factors, a key input for decision making (Bormann et al. 1994; Jensen, Bourgeron, Hessburg, et al. 1994), it is precisely in the realm of decision making that the relationship between humans and nature is most clearly distinguished. Particularly for public land management agencies in the United States, the decision making process is largely constrained by matters of policy and law. The National Environmental Policy Act, Organic Act, Multiple Use Sustained Yield Act, Endangered Species Act, National Forest Management Act, Clean Air Act, Clean Water Act, Federal Advisory Committee Act, and others, can all be seen as defining the planning and ultimate decision making which will occur under the EM framework (Jensen and Everett 1994; Keiter 1988; Scientific Integration Team 1994). An EM approach perhaps offers a more functional way to incorporate better data regarding the natural environment into decision making, but EM does not in itself alter the greater political decision making process as prescribed by law and policy.

Perhaps more to the point, literature regarding the EM framework tends to distinguish between decision making inputs, including ecological, social, economic, and technological assessments, from the actual decision making process (Bormann et al. 1994; Kaufmann et al. 1994; Scientific Integration Team 1994). As for actually choosing specific management scenarios within EM, a suggested selection criteria includes various social and economic assessments, as well as ecological assessment of what a sustainable flow of resources might be, but it does not include assessment of ecological parameters outside of the sustainability realm (Bormann et al. 1994; Everett, Hessburg, et al. 1994). Decision making itself is conceived of as a "sociopolitical and values" process (Grumbine 1994), and a process of "collaborative negotiation" (Daniels et al. 1994), in which "societal approval" (Bormann et al. 1994) is a key factor.

In sum, while the EM framework places great importance on an understanding of natural processes, natural variability, natural disturbance, natural diversity, and the role of native species as necessary for sound planning and management, and may even value ecosystems themselves as "objects of respect and admiration" (Goldstein 1992), it does not change the "human dimension" (Human Dimension Study Group 1994; Kaufmann et al. 1994) of management; the basic reality that "management" of natural resources, by definition, involves imposing change upon ecosystems in order to derive "goods and services" for human benefit.

On the other hand, an emphasis on EM's role as a management system for commodity production from natural resources does not completely overshadow EM's explicit recognition that there are limits to ecological capacity (Scientific Integration Team 1994) or resiliency (Kaufmann et al. 1994) and hence limits to the sustainability of resource exploitation (Kaufmann et al. 1994; McConnell et al. 1994; Salwasser et al. 1993). Thus, while the human element of EM may overtly dominate the decision making process, elements of ecological science, and in particular those concepts related to naturalness which have been reviewed at length above, nevertheless remain a crucial factor as well. In this sense, the two elements, although elaborately distinguished in most descriptions of the EM framework, are at another level inseparable, and the struggle to delineate a relationship between the human and natural realms might itself be better viewed as part of the EM process, rather than an obstacle to its implementation.

The preceding sections of this literature review have focused on philosophical, cultural, ideological, and socially normative perceptions of nature and naturalness. More specifically, the review of literature pertinent to an Ecosystem Management framework directly addressed the ways in which concepts of naturalness are defined and employed within a more scientifically oriented, ecological realm. These are all important factors in building an overall understanding of the conceptualization of naturalness in both a general and an ecological or resource management context (Merchant 1993; Soper 1995; Worster 1993); however, none of the literature so far discussed has dealt with explicit conceptual systems or frameworks for linking these cultural and ecological perspectives. And it is only through such frameworks that there is hope of forming some coherent way of incorporating concepts of naturalness and nature into the realm of resource management.

As has been discussed at length, the definition of naturalness in an ecological context is, at the least, controversial and problematic. However, while simple answers may seem elusive, the question of what can be considered natural is nevertheless a kind of "fundamental question" confronting scientists and resource managers alike; in other words, it is a question which underlies basic assumptions of scientific understanding and theory, and hence policy and practice as well (Pickett, Kolasa, and Jones 1994).

In this regard, it has been pointed out that there is a difference between understanding the way concepts of nature are socially influenced and defined, versus attempting an understanding of the actual content and structure of what we call nature itself (Lease 1995). Despite the importance of recognizing how the influence of ideologies and theoretical constructs shape perceptions of what naturalness entails (McLaughlin 1993; Soper 1995), and how it may be scientifically studied (Pickett, Kolasa, and Jones 1994), it is, at some fundamental level, also important to avoid conflating social norms about what is natural with the concrete characteristics of natural things themselves (Evernden 1992).

Science in general, and ecology in particular, can be seen as dealing with two kinds of subjects; "natural" ones and "formal" or "theoretical" ones (Kiester 1980; Louie 1985). For the most part, science is seen as focusing on the need to "reduce" specific "natural kinds," through abstraction, to more generalized "kinds" or categories. In essence, the only way for science to study naturalness in through the reduction of "nature" to a "description of nature" (Heisenberg 1958). This can be accomplished by isolating key characteristics of naturalness, developing a series of discrete descriptive categories based on those characteristics, and then developing a framework within which those categories can be related to the individual phenomena which seem to comprise nature itself (Heisenberg 1958).

Within the realm of resource management, a number of such frameworks have been put forward for understanding the relationship between humanity and its natural environment. As discussed previously, various ideological stances may be taken in this regard, each of which may be in varying degrees either predominantly nature-skeptical versus nature-endorsing (Soper 1995).

Additionally, Soper (1995) has suggested that application of concepts of nature and naturalness within an ecological context may take three basic approaches: in what may be termed a "metaphysical" concept, through philosophical or ideological stances in which humanity and nature are either seen as antithetical or indistinguishable; in what may be termed a "realist" concept, in which the structures and processes that form the basis for scientific study are all constituents of "nature"; and, in matters concerning human society, a "lay" concept of nature, in which "naturalness" is used to contrast urbanized, domesticated, or industrialized environments.

Attempts to develop formal, idealized frameworks for understanding naturalness, using theoretical categories, can on the one hand be seen as attempts to isolate technical or "realist" interpretations of naturalness from more loosely defined "metaphysical" or "lay" interpretations (Soper 1995). On the other hand, for the purposes of managing natural resources, it seems that some fusion between the "metaphysical," "realist," and "lay" concepts is necessary, to the extent that, in terms of applying scientific methods to managing natural resources for the benefit of society, it simply proves impossible to isolate philosophical, scientific and social aspects of a common concept (Merchant 1989; Pickett, Kolasa, and Jones 1994; Worster 1993).

There are of course, those who simply maintain that there can be no distinction between the human/cultural realm and the natural realm, and any discussion of frameworks for understanding the relationship between the human and natural realms must begin by acknowledging this metaphysical or ideological stance. Such arguments generally take one of two basic approaches relevant to resource management.

First, there are those who simply reject, on ideological grounds, the possibility that there can be anything "natural" separate from human manipulation or outside of human perception. This perhaps represents a purely nature-skeptical and metaphysical approach, but it is often specifically voiced within an ecological or scientific context (Koromondy 1974; Pollowy 1994; Rassler 1994; Scherer 1994). Second, there are those who maintain that, while such a distinction may have had reality in pre-technological times, these kinds of natural conditions ceased to exist thousands of years ago, with the advent of hunting and agricultural techniques.

Most frequently, this second argument is made in regard to European environments (Bradshaw 1994; Oxford 1994), but it is also at times applied universally, pointing to the at least subtle effects of human influence in even the most remote parts of the earth (McKibben 1989; Williams 1993). This second approach may be seen as an attempt to encompass both a nature-skeptical metaphysical approach with a "realist" approach, accepting the theoretical need to acknowledge some definable realm for the natural sciences, but still rejecting the reality of an independent nature on ideological grounds.

Both of these perspectives are in opposition to the paradigms which will be discussed in the following paragraphs and, while perhaps representing a minority view, certainly have a substantial number of adherents. However, they are also far from universally accepted: by way of example, ecologists working in heavily developed and manipulated parts of Europe such as Germany (Sukopp 1976), Poland (Oleksyn and Reich 1994), Britain (Newbold 1989), and France (Aronson et al. 1993), as well as cultural geographers (Jackson 1984), human ecologists (Clapham 1981; White and Renner 1936), and land-use planners (Doxiadis 1977; McHarg 1993) all have been quite comfortable with discussing as "real" nature and a quality of naturalness.

Taking a "realist" perspective, an at least partly nature-endorsing position seems based in one simple observation; that is, while there may indeed be no part of the earth which is uninfluenced by human cultural practices, neither is there any part of the earth which is purely a cultural artifact (Van der Maarel 1976). From this observation, it is argued that, although it may often prove difficult to distinguish natural and cultural aspects of environments, this does not in itself preclude such distinctions, nor lessen the importance of recognizing the respective roles of natural and cultural processes in seeking ecological understanding (Castilla 1993; Pickett and McDonnell 1993).

The simplest approach to distinguishing between natural and non-natural aspects of environments in by contrasting "natural" things with some binary opposite (Grange 1992; Tuan 1974). In this context, a variety of terms have been used to serve as the opposite of natural. For example: natural versus cultural (White and Renner 1936), natural versus built (Kaplan and Kaplan 1989), natural versus manipulated (Simmons 1982), natural versus artificial (Hausman 1975), natural versus semi-natural (Newbold 1989), or natural versus anthropogenic (Coleman and Hendrix 1988).

Other pairs of terms may also be employed in a similar and yet broader context, such as authentic versus inauthentic (Relph 1976; Tuan 1974) or, in more metaphysical terms, subjective versus objective (Harvey 1989). In each case, what is attempted is delineation of simple, contrasting terms to enable a coherent dialogue concerning these differing aspects of environments.

Likewise, numerous authors have attempted to differentiate between the natural and human/cultural realms by contrasting the natural "biosphere" with a "noösphere," the world as dominated by the "mind of man," (Barrett 1981; Cain 1966), an "anthroposphere,"

the world as dominated by human artifacts (Doxiadis 1977; Vink and Davidson 1983), or a "technosphere," the world as altered by human technology (Morowitz 1991; Naveh and Lieberman 1984). In each of these models, there seems to run the common theme that the "human environment" is a subsystem of the greater biosphere (Maldonado 1972), and that, while there are both obvious and obscure linkages between these natural and culturally derived ecosystems or ecosystem components, they are nonetheless distinguishable (Barrett 1981; Clarke 1986).

However, it is also plain that, while such binary terms are conceptually useful, the complex and often subtle interactions between natural and cultural elements in most ecosystems makes such clear distinctions difficult and largely unworkable for specific scientific of managerial applications (Noss and Harris 1986; Pickett and McDonnell 1993). For this reason, it seems more useful to construct a framework which looks at varying degrees of naturalness and cultural influence (Sukopp 1976; Van der Maarel 1976), perhaps best expressed in terms of a gradient (McDonnell and Pickett 1990; Van der Maarel 1976) or continuum (Anderson 1991; Soper 1995; Taylor 1986).

A number of such gradient or continuum schemes have been proposed; several of these are displayed in table 4. Although each of the schemes displayed in table 4 differ at least slightly in their approach, the common intent is to establish some standard categories while allowing for gradations along a continuum between them.

In the case of Doxiadis (1977), the proposed scheme was meant to serve as a kind of taxonomy for land allocation within a far more detailed planning framework. Bakker (1979), Van der Maarel (1975), and Odum (1983) each framed their proposals within a similar land-use context, although with a more directly ecological orientation, while the three categories used by Johnson and Carothers (1987) were employed in a more specifically ecological-research context. Each apparently attempts to define categories which can serve as benchmarks along the proposed continuum, while recognizing that there are no truly discrete categories as such.

Continua of Naturalness and Categories

Reference

Natural

Near-

Natural

Semi-Natural

Agricultural

Near-Cultural

Cultural

Van der Maarel 1975

Natural

Subnatural

Seminatural

Agricultural

Bakker 1979

Naturareas

Cultivareas

Anthropoareas

Industrareas

Doxiadis 1977

Natural

Domesticated

Fabricated

Odum 1983

Natural

Naturalized

Exotic

Johnson and Carothers 1987

The continua of naturalness described above provide a framework for understanding the relationship of natural and cultural aspects of areas or systems; however, they still leave open the basic question of exactly how to distinguish a character of naturalness from one of culturalness. Several basic approaches have been used in this regard. One approach, although not amenable to the concept of naturalness occurring as a continuum, is to regard only those systems or areas which are "pristine" in some absolute sense as natural (Bakker 1979; Oxford 1994), while another approach is to consider naturalness as ceasing with human cultural influence beyond the Paleolithic level (Bradshaw 1994).

Other approaches consider humans themselves to be natural, but not the effects of cultivation and construction which accompany settlement (Doxiadis 1977), the effects of technological societies (Wagner and Kay 1993), or the effects of industrialized cultures (Hayes, Riskind, and Pace 1987; Kilgore 1987). In North America, it is often suggested that the advent of non-indigenous European cultures marks a significant benchmark which can delineate the advent of non-natural cultural effects (Bonnicksen and Stone 1985; Johnson et al. 1994).

Others, however, have pointed out that these strategies all mix ecological and social perspectives on what constitutes naturalness, and suggest that, in ecological terms, it is more appropriate to focus only on ecosystem function (Coleman and Hendrix 1988; Lease 1995). From this perspective arises a distinction between inherent, and thus "natural" ecosystem processes, and those which are subsidized by humanly organized input (Anderson 1991; Coleman and Hendrix 1988; Odum 1983).

Although the pre-settlement (or in North American pre-European settlement) benchmark is most commonly employed by resource managers, particularly within the EM approach (Swanson et al. 1994), this more functional approach of focusing on ecosystem processes and their subsidy or lack thereof is arguably more appropriate within an ecological context (Anderson 1991).

There have been a number of proposed methods for evaluating natural areas or systems from this more functional perspective, most focusing on a particular aspect of the ecosystem of interest and using resulting data to make inferences about overall ecosystem states. For example, there has been the approach of "rapid assessments," quick surveys used to evaluate areas under consideration as nature preserves, which focus mainly on species composition in terms of indicator species, such as large mammals, birds, and trees (Abate 1992).

In a similar fashion, surveys of biodiversity have also been suggested as a way of indicating the overall condition of system processes (Handel et al. 1994; Noss and Harris 1986). Others have used more detailed surveys of species composition, incorporating consideration of native/non-native status (Johnson and Carothers 1987) and species rarity (Keddy and Sharp 1994), or expanded such studies to evaluating community structure (Karr 1987; Peterman 1980) or community trophic-flow patterns (Havens 1993).

An even more closely function-based approach has also been employed by directly focusing on ecosystem processes, especially those related to succession (Connell and Slatyer 1977; McIntosh 1980) or disturbance (White and Pickett 1985). Studies concerned directly with these aspects of ecosystem process may in turn focus on either description of disturbance regimes themselves (Bonnicksen and Stone 1985; Christensen 1988) or in studying the response of ecosystems to such disturbance (Goudie 1986; Westman 1991).

Other factors which have been proposed as the basis for evaluating natural systems have included consideration of ecological "uniqueness," an attempt at estimating the overall importance of a species or community for greater ecosystem function (Orians and Kunin 1990), or similarly, consideration of long-term evolutionary dynamics (Erwin 1991).

These criteria have also been variously combined into indices used to evaluate ecosystems. In this vein, a simple formula for evaluating natural areas was proposed by Sankovskii (1991), using four factors: relative fragility, uniqueness, size of area, and intensity of destructive factors. Others have developed far more elaborate indices, using "suites" of indicators (Kelly and Harwell 1990), often describing ecosystem states in terms of "health" (Schaeffer, Herricks, and Kerster 1988) or "integrity" (Karr 1987; Munn 1993).

Although these concepts of health and integrity are related to that of naturalness (Kay 1993), a review of pertinent literature finds few attempts at specifically developing quantifiable approaches to evaluating naturalness itself. However, those attempts that have been made also take a broadly based functional and ecological approach.

For example, Sukopp (1976) developed an extensive matrix for evaluating "degrees of naturalness," including a detailed characterization of six classes of land-type along a spectrum from "natural" to "devastated." Sukopp's framework was then used by Van der Maarel (1976) to develop a "naturalness index" based in six evaluative criteria: hemerobiotic state (relative state of cultivation versus biotic integrity), changes in site substrate, changes in vegetation structure, changes in floristic composition, loss of native species, and gain of non-native species.

More recently, Anderson (1991) proposed that the naturalness of an area could be quantified with a combination of three indices: (1) the degree to which the system would change if humans were removed, (2) the amount of cultural energy used to subsidize ecosystem function, (3) the complement of native species present compared with the suite of species in the area prior to human settlement. Similar, yet more detailed approaches, have also been developed in efforts to evaluate the project work of restoration ecology, focusing specifically on quantifying the dynamic functional attributes of ecosystems (Berger 1990; Pacific Estuarine Research Laboratory 1990).

While the concepts discussed above, developing a continuum delineating relative degrees of naturalness and evaluating naturalness using ecologically-based criteria or indices, are in themselves useful approaches to ecological study, they cannot be effectively applied to solving problems of resource management unless they are considered within a broader planning framework.

Several such planning models have in fact been proposed, some of which are tied directly to these concepts. For example, Van der Maarel (1976), in conjunction with his proposed continuum of degrees of naturalness, developed a "physical planning model," differentiating between social, economic, administrative, ecological, and spatial systems, all of which are seen as interacting as a greater system through direct-function and feedback relationships. Likewise, Eugene Odum (1983) developed a set of "compartment models" for environmental-use planning, based in distinct categories of natural, domesticated, and fabricated environments, which were then incorporated within a fairly complex management system for allocating land use.

Similar planning schemes have also been put forth framed in the context of the inter-relationships described by the biosphere and noösphere concepts (Cain 1966), or in terms of delineating the relationships between natural environments and socially driven planning (Naveh and Lieberman 1984; Vink and Davidson 1983).

Landscape ecologists have in particular given attention to developing these kinds of planning models and recognizing the distinct but firmly linked cultural and natural aspects of environments. One such model, originally proposed by Van der Maarel (Vink and Davidson 1983) especially emphasizes the concept of naturalness as forming a true continuum, and one which is meant to be incorporated within a functional management system, not merely as a scale to rate the condition of reserves.

Likewise, models suggested by Naveh and Lieberman (1984), and Boyden (1993) demonstrate that overall intensive land-use planning is served by the distinction between degrees of naturalness. The model of Naveh and Lieberman is of particular interest in that it explicitly attempts to define relationships between the biosphere, the culturally imposed "technosphere," and a "total human ecosystem" encompassing both of these. Naveh and Lieberman's model is also of note in that it incorporate as well the concepts of energy subsidy and artifact construction in defining these relationships.

Similarly, Boyden (1993) attempts to delineate the functional relationships between social, cultural, biotic, and natural aspects of the human ecosystem, using a systems approach to distinguish between the physical components of both nature and society, and their functional interactions. Boyden's model is also of interest in that he differentiates the "biophysical actualities" of nature and culture from the "abstract culture" which, although intangible, is yet seen as a key factor in how humans interact with nature.

These and other similar planning models seem to reflect an awareness on the part of land-use planners that effective planning must incorporate information supplied through ecology and an ecosystem paradigm, as well as an explicit focus on the relationship between natural ecosystems and human cultural systems. In the 1960's, planners such as Ian McHarg suggested that an

understanding of ecosystem composition and natural processes was the necessary basis of effective land use planning (Belknap et al. 1967; McHarg 1969), while ecologists called for applying ecological science in a "preventative and regulatory" manner (McHale 1970).

Similar opinions became voiced by resource managers as well, (Holling 1978; Walters 1986), leading to managerial frameworks which incorporated planning models similar to those proposed by Van der Maarel and others (Salwasser 1986; Vink and Davidson 1983). Specifically within resource management, these approaches have recently received renewed scrutiny and a measure of acceptance, particularly within the framework of Ecosystem Management (Bormann et al. 1994; Kaufmann et al. 1994; Scientific Integration Team 1994). However, they seem to remain poorly incorporated within the realms of resource allocation and decision making, and this situation seems in no small part due to the persistent problem of naturalness remaining a concept diversely interpreted and defined.

Although the frameworks described above offer some perspective on how resource managers can incorporate concepts of naturalness within their disciplines, none of these frameworks can be regarded as widely accepted or employed. Perhaps more significantly, they still leave open to debate exactly how the quality of naturalness should be described and evaluated. Proposals for developing indices or quantifiable measures of naturalness, such as Van der Maarel's (1976) or Anderson's (1991), may represent important contributions toward developing some standard methodology in this regard, but there certainly appears to be no accepted approach for doing so.

Like any other group of human beings, the people involved in resource management come to their field variously influenced by the many philosophical, historical, and ideological schools of thought which have been discussed previously. Kellert (1995, 1996) attempted to discover what kinds of attitudes were held toward nature in a normative sense, and although he found some clear patterns of attitude, he also found these attitudes to be complex and somewhat self-conflicting.

Reviewing the literature of ecology and resource management, traces of these conflicting philosophical and ideological approaches seem evident as well; but the literature does not seem to directly address how those involved in resource management, as a group, define the characteristic of naturalness, nor whether it is perceived that some standard or normative definition of the concept exists.

Given the extensive references made to the term "natural" within the discourse of both ecological science and resource management, particularly within the EM management framework, and given that naturalness seems largely an implicitly rather than explicitly defined concept, an effort at direct assessment of opinion seems the only way to clarify how the concept of naturalness is defined and used. It is in light of this remaining ambiguity and the need to clarify concepts of naturalness specifically as held by those involved in resource management that the case study portion of this project was developed.