The Multifaceted Aspects of Ecosystem Integrity

• Abstract
• Introduction
• Ecosystem Integrity
• Why Preserve Biodiversity?
• Assigning Economic Value
• Managing and Mismanaging Ecosystems
• Conclusion
• Responses to this Article
• Literature Cited

ABSTRACT

KEY WORDS: adaptive management; biodiversity; complexity and stability; conservation strategies; disturbance, anthropogenic; disturbance, natural; ecosystem integrity; ecosystem functioning; ecosytem structure; natural resource management; resilience; sustainable development.

INTRODUCTION

The physician's task is to evaluate and maintain healthy functioning of an individual; the environmental manager's, to evaluate and maintain healthy functioning of an ecosystem. The analogy is seductive, and has led to the development of the concept of ecosystem health (e.g., Costanza et al. 1992), which presupposes defining a normative state of natural systems and identifying limits of human intervention. Obviously, this notion is appealing. A healthy body is physically vigorous and free from disease. Similarly, a healthy ecosystem or community might be indicated by ranges of values considered to be normal, and by attributes that are regarded as stable and sustainable, whereas pathological conditions are indicated by the opposite (Schaeffer et al. 1988, Ryder 1990, Costanza et al. 1992, Rapport 1992, Freedman 1995). To large extent, however, this concept is rooted in the organismic theory of ecology promoted by Frederick Clements at the beginning of this century, and is based on the idea that biological communities are structurally and functionally like organisms. In the Clementsian view, communities are recognized as having their own identities that may change in time but eventually reach a fixed, normative balance state at the end of the successional process. This view of ecosystems thus directs attention to development of equilibrium theories that ignore dynamic features. Moreover, the "superorganism"paradigm ignores the degree to which ecological communities are open, loosely defined assemblages with only weak evolutionary relationship to one another (Levin 1992). This definition is highly dependent on the scale of description, and, GAIA notwithstanding, the ecological community is not an integral evolutionary unit in the sense that an individual is. Here the metaphor fails.

Furthermore, the emphasis put on stability, unique equilibria, and normative states has historically promoted a view of a "benign Nature" able to cope with any sort of anthropogenic interference and manipulation, because trials (and errors) of any kind can be made with the assurance that recovery is always possible once the source of disturbance is removed (Holling 1987). In reality, all natural ecological systems change over time, and it is extremely difficult to determine a normal state for communities whose measurable properties are often in flux, either because of natural disturbance or because of internal ecological mechanisms (Ehrenfeld 1993). Ecosystems are seldom close to equilibrium, a fact not recognized adequately in many environmental assessments (Reice 1994). Some communities may be consistently affected by high levels of localized disturbance that allow for the coexistence of species competing for the same limited resource (e.g., intertidal communities), whereas others may experience periodic catastrophic episodes, such as floods, forest fires, or pest outbreaks, that reset the timing of successional cycles. Furthermore, disturbance may occur on a wide range of temporal and spatial scales (Levin 1995). Therefore, in defining ecosystem health, it may be difficult to separate the effects of human and natural disturbance. To confuse matters more, the components of an ecosystem may be mutually connected in a variety of ways and, thus, may exhibit an ensemble of different functions. Hence, any attempt to evaluate ecosystem health will depend upon which functions and which components of the ecosystem we are considering.

Because different communities may exhibit completely different spatial and temporal patterns of species abundance and of functional activity (Levin 1995), the attempt to redefine the notion of health for any particular system can weaken any larger, unifying idea of health. Most importantly, the organismic theory of ecology, on which the notion of ecosystem health is grounded, fails to recognize that ecosystems are not uniquely identified entities, nor are they defined by sharp boundaries (Karr 1994). Instead, they are loosely defined assemblages that exhibit characteristic patterns on a range of scales of time, space, and organization complexity. A more promising approach to ecosystem management is to recognize that various genetic, competitive, and behavioral processes (rather than states) are responsible for maintaining the key features of observed ecosystems, and that the dynamics of these processes vary with the scale of description. Holling (1992) proposes a dynamic view of ecosystem cycling through a spiraling developmental path, characterized by different phases. Here, the emphasis is on variability, spatial heterogeneity, and nonlinear causation.

Problems aside, it is essential that we develop measures of ecosystem functioning, for which we prefer the word "integrity." The notion of integrity must accept the dynamic view incorporating processes. It must recognize a human perspective, the ability of an ecosystem to continue to provide the services that humans expect. For managed ecosystems, the ability to supply products such as food or timber may provide the integration; for natural systems, other valuations will enter. It is important to recognize that these are imposed measures, conditional on a definition of "use" for a system. In this way, the notion of integrity differs fundamentally from a unique definition of "health" as an evolved aspect. It is a tool for management.

Ecosystem integrity is so complex an issue that a single indicator or operational definition is insufficient to grasp its multifaceted aspects. The aim of this paper is to review carefully the concept of ecosystem health and integrity; to identify links among biodiversity, ecosystem functioning, and resilience; and to stress and distinguish the effects of natural and anthropogenic disturbance. Throughout, we direct attention not only to theoretical aspects, but also to links between ecosystem theory and management, and to practical implications based on different notions of ecosystem integrity.

ECOSYSTEM INTEGRITY: theoretical aspects and practical implications

  1) In resource-based systems, such as forests, fisheries, and croplands, decision makers are inclined to devote considerable effort to keeping the system within desirable stability domains that guarantee optimal exploitation rates. They attempt to prevent any shift toward a new mode of functioning, by reducing environmental stochasticity that could push the system away from its optimal state and cause undesirable economic or ecological effects. This imposed resiliency reduces the sensitivity of the system to exogenous factors that might adversely affect the exploited population(s), but at the cost of sacrificing information about the dynamic properties of ecosystems in changing environments.

  2) Conversely, it is the dynamic processes themselves that guarantee the functioning of an ecosystem, and any successful effort to constrain natural variability will eventually lead to self-simplification and fragility. Therefore, it is neither desirable nor reasonable to eliminate the natural successional cycle of a natural system. Keeping one population at a constant level may lead to changes in other related species, so that features of the biophysical environment that usually have been perceived as constant begin to change and to produce a system that is structurally different from the original one. The resilience of the system to change is embedded in its heterogeneity and dynamic properties, and especially in these hidden variables.

Although they recognize the complex, dynamic nature of ecosystems, resource managers, therefore, often aim to achieve constancy through externally imposed regulations, attempting to reduce the probability of events that are perceived as ecologically or economically undesirable, such as floods, pest outbreaks, and fire (see Appendix 1 ). Unfortunately, policies to eradicate these natural nuisances have frequently led to dramatic impacts on communities. Numerous illuminating examples are described in Holling (1987), with reference to other managed systems, such as the salmon fisheries in North America; the conversion of the semiarid savanna ecosystem to productive cattle grazing lands in Africa, the United States, India, and Australia; and the malaria eradication programs in developing countries. All these examples share the common feature that a blind attempt to control some "undesired" ecological or economical effect has been successful only over a short span of time. However, the price paid to achieve this short-term objective has been a qualitative change in the behavior of the system; the alteration of natural, long-term cycles has led to a loss of resiliency and has produced system crises much larger that those occurring in unmanaged ecosystems.

Pickett and White (1985) state, "An essential paradox of wilderness conservation is that we seek to preserve what must change." On the other end, if ecosystems experience fluctuations and changes generated by internal ecological mechanisms, this does not mean that any change should be accepted. In contrast, Botkin (1990) states, "we must focus our attention on the rates at which changes occur, understanding that certain changes are natural, desirable, and acceptable, while others are not."

Toward a paradigm for ecosystem integrity

The concept of ecosystem integrity is not free from criticisms (Anderson 1991, Rolston 1994). However, rather than engaging in endless debates over which is the best and most comprehensive definition of integrity, we agree with Noss (1995a) that it is much more useful to characterize in detail the functional and structural aspects of ecosystems to provide a conceptual framework for assessing the impact of human activity on biological systems and to identify practical consequences stemming from this framework.

Reductionism vs. holism

  1) A reductionist approach emphasizes the structural aspects of natural systems and focuses on individual species and population dynamics of species within isolated ecosystems (Soulé 1986);

  2) A holistic approach focuses on macro-level functional aspects (in particular, energy flows, nutrient recycling, and productivity), to some extent neglecting historical and evolutionary factors and ignoring most of the details observed at smaller scales of functional organization and of the spatial and temporal distribution of organisms.

Of course, the structural and functional perspectives on biological systems are not mutually exclusive. They simply reflect broad streams in which many branches of ecological theories have flourished. In our context, they are useful in outlining two rather different approaches to the compelling issue of preserving ecosystem integrity (King 1993). In the extreme, in fact, they lead to very different definitions of integrity:

  1) Strict attention to the structural aspects of ecosystems, as represented primarily in species composition, leads to a definition in which the loss of even one species or the damage of a link between some components implies a loss of integrity, because the ecosystem is no longer "complete" or "whole."

  2) On the contrary, from the perspective of functional integrity, redundancies within functional groups make the biological composition less relevant.

There is merit in either definition of integrity, or in their combination; the relevance of either depends on the perspective of the investigator and on the way ecosystem services, resource species, aesthetic values, and other aspects are balanced (Levin 1997). Furthermore, structure and function are linked. Many macroscopic properties (primary productivity, in particular) are very resilient to changes in system structure, at least on short time scales. Even when subject to high levels of disturbance (and, thereby, to substantial changes in their structure), a system may be able to preserve its macro-level functions, such as primary productivity, and some macro-level indicators may not show any appreciable change.

An often-cited case is that of the American chestnut (Castanea dentata). This canopy species, once fairly dominant in the deciduous forest of eastern North America, has been wiped out by the introduced blight fungus Endothia parasitica. The population has been replaced by other shade-tolerant species (Hepting 1971, Spurr and Barnes 1980), without substantial changes in the primary productivity of the forest. Estuarine communities provide an example of biological systems subject to a high level of disturbance, such as hurricanes and floods, and characterized by a high variability in community composition over time. Yet some macro-level functions are amazingly resilient to alteration in structure: productivity rapidly recovers after major catastrophic events that completely reset the biological clock of the community (Costanza et al. 1993).

According to a strict interpretation of the functional approach to ecosystem integrity (King 1993), a change in ecosystem structure that does not appreciably change the qualitative and quantitative functional aspects should be interpreted, at most, as a minor loss of integrity. What is missed in this view, however, is that loss of diversity within functional groups may weaken the ability of the system to adapt to catastrophic changes on longer time scales.

Neither pure structuralism nor pure functionalism is completely appropriate; both contain validity, but within some boundaries. Putting equal emphasis on every piece of biodiversity is ecologically unsound and tactically unachievable (Walker 1992). On the other hand, the assessment of ecosystem integrity based only on a few macro-level indicators, such as primary productivity or other measures of energy and matter flows, may obscure other ecosystem properties that ultimately determine the resilience and stability of the ecosystem to several sources of disturbance. Indeed, the two extremes simply represent boundary points in a multidimensional continuum, in which a variety of measures of differing levels of detail may be applied.

In the following sections, we will briefly sketch some of the most widespread theories about the importance of species diversity and the relationship between biodiversity and ecosystem properties.

WHY PRESERVE BIODIVERSITY? Linking Biodiversity to Ecosystem Functions

Keystone species and functional groups

In most cases, it is indeed groups of species, rather than individual species, that assume importance, forming "keystone groups" or "functional groups," a generalization of the notion of keystone species. Functional groups (guilds) are collection of species that perform the same functions and that, to some extent, may be substitutable and viewed as a unit (Schulze 1982, Solbrig 1994). For example, removal of a numerically dominant species may result in its replacement by functionally similar competitors that had been suppressed, leaving untouched macro-level indicators of ecosystem functioning (like productivity, or the amount of matter processed). Yet, loss of species within a guild may reduce the long-term resilience properties of the system, and may lead to noticeable change in short-term system dynamics (Levin 1997).

The role of ecological redundancy

Communities viewed in terms of functional groupings, in general, prove to be much more stable and predictable than when viewed in terms of species composition (Hay 1994); this is largely a property of averaging. For example, convergent biogeographical patterns in ecosystem organization have been clearly discerned when distinct species of subtidal algal communities in Maine, Washington State, and the Caribbean have been grouped based on common morphological attributes (Steneck and Dethier 1994). In fact, changes in species populations within a functional group usually occur on a much faster time scale than dynamics among groups, allowing a hierarchical decomposition of the system dynamics (Simon and Ando 1961, Iwasa et al. 1987, 1989). Therefore, on the short time scale, a reduction of within-group heterogeneity is not likely to change ecosystem properties appreciably, and biotic detail is probably irrelevant. On the slightly longer scales, biotic diversity and consequent feedbacks may fundamentally alter the responses of systems to stress. Determining precisely what is "short" or "slightly longer" is, of course, very subjective. However, Bolker et al. (1995), modeling a forest in which shifts and feedbacks due to natural selection occur within and among functional groups, found that significant responses of the biota can be expected on time scales of 50-100 years, a relatively short time for forests and for the horizon of interest to humans.

Complexity and stability

A central question

Disturbance and temporal scale of investigation

Natural and anthropogenic disturbance

Ecosystems affected by human activity usually exhibit reduced resistance to natural stresses, such as fires, droughts, pests, and diseases, in a positive feedback fashion. When a native grassland is converted to a monoculture of corn, for instance, the resulting system (which is highly simplified with respect to the original one) is inherently unstable and needs huge inputs of energy and material, such as fertilizers and pesticides, to remain in the desired condition (Noss 1995a). On the other hand, human disturbances that mimic or simulate natural disturbances are less likely to threaten ecological integrity than are disturbances radically different from the natural regime. Therefore, in managing ecosystems, the goal should not be to eliminate all forms of disturbance, but rather to maintain processes within limits or ranges of variation that may be considered natural, historic, or acceptable (Noss 1995a).

ASSIGNING ECONOMIC VALUE TO BIODIVERSITY

To conservation biologists, putting an economic value on biodiversity may seem both arrogant and useless. Aldo Leopold (1953) wrote: "The last word in ignorance is the man who says of an animal or plant: What good is it ? ... To keep every cog and wheel is the first precaution of intelligent tinkering." Placing an economic value on biodiversity prevents us from coping with the root causes of loss of diversity and from recognizing the value of the environment, other than as a commodity to be exploited (Meyer and Helfman 1993); it forces us to accept the old economic paradigm that assumes a perfect substitution between the natural capital and market capital, and it reinforces the technological premise that makes the biological impoverishment of the planet inevitable (Ehrenfeld 1988). Given the present interest rate on money, there is no hope that traditional economic approaches will preserve natural systems. Clark (1973a) elegantly proved the fallacy of this approach, demonstrating that it could be economically preferable to kill every blue whale left in the ocean and to reinvest the profits in the stock market, rather than waiting for the species to recover to the point at which it could sustain an annual catch. An economist might argue that the problem, in this example, is an inadequate evaluation of the resource; if aesthetic values are important, simply put a price on them. In practice, however, this ignores the complexity of valuation, including the problem of who speaks for the less powerful or for future generations.

Leaving apart the fact that many, if not most, species do not seem to have any conventional value, even a hidden one (Ehrenfeld 1988), it should be borne in mind that the true economic value of any piece of biodiversity cannot be determined without considering the value of biodiversity in the aggregate. Classification of diversity by species is only one possibility, and other forms of diversity, should be taken into account during the evaluation process (see Appendix 3 ). The patchiness and diversity of ecosystems across the landscape (forest, wetland, grassland, estuarine, and marine ecosystems) play a crucial role in retaining soil and nutrient availability and purifying the air and the water.

In conclusion, assigning an economic value to species is extremely complicated, highly subjective, and very simplistic. It is difficult to attach levels of probability of potential benefits in the absence of appropriate information. The process of assigning values, especially ones that reflect potential future use, is extremely subjective and is inevitably affected by the present poor scientific knowledge of ecosystem structure and functioning. Instead of arrogating to themselves the valuation of species, scientists should limit their advice to science, informing the decision-making process, but leaving complex trade-offs to be resolved by the stakeholders. This is a highly controversial area, of course, and it is essential to find new mechanisms for inserting science into the decision-making process. Multisector partnerships involving science, industry, the public, and decision makers offer great hope for the future.

MANAGING AND MISMANAGING ECOSYSTEMS

This subject has been so thoroughly studied, in theory and practice, that it is certainly one of the best understood areas of environmental science. Unfortunately, there are very few ecosystems in which sustained exploitation has proved to be successful (Clark, 1973b, Ludwig et al. 1993). All over the world, renewable resources have been systematically overexploited and diminished to local decimation or extinction. It has been argued that the causes of mismanagement are rooted in the inherent complexity of biological communities and in the high level of environmental stochasticity that confounds any effort toward an ecologically sound exploitation practice. On this point, one of the most attractive notions is that of implementing adaptive methods of management (Holling 1978, Walters 1986) that focus on the links and the mutual interactions among biological factors, economic considerations, and natural variability. The use of modern decision theory (Berger 1985, Lindley 1985, Mangel 1985, Hilborn 1987) allows natural resource managers to include environmental uncertainty explicitly in the decision process, by continuously monitoring the managed system and recursively upgrading management protocols in response to observations. Yet, with some exceptions, adaptive management remains more a hope that a realization.

Resource management, even when included in an adaptive framework, traditionally has focused on relatively small scales of spatial and functional organization, ignoring the broader ecological and environmental context in which exploited resources are often, if not always, embedded. Second-order, indirect, and sometimes irreversible impacts have been systematically neglected in the effort to mitigate first-order direct impacts on target resources, and to increase production and possibly reduce fluctuations in the exploited biomass. Consequently, resource exploitation may be sustainable with respect to the commercial species, but may seriously threaten the viability of other components of the ecosystem. For instance, Mangel et al. (1993) reported that the current worldwide catch of penaeid shrimp might be sustainable, yet an estimated 89-90% of shrimp trawlers' catches are nontarget species. One of the nontarget species most affected by shrimp exploitation is Kemp's Ridley turtle (Lepidochelys kempi), which has experienced a 99% decline in population size in the last 50 years (Pritchard 1990); several other species are accidentally caught (Andrew and Pepperell 1992), although their decline is far less likely to be noticed.

Focusing strictly on target or endangered species is rarely the best strategy: the species-by-species approach may be inefficient for several reasons (Carrol et al. 1996). First, it may be difficult to advocate the importance or need for a particular piece of biodiversity, as long the relative importance of specific species to the overall functioning of the ecosystem is not known. Second, conservation measures may be activated only when most of the damage has been done already, and the detrimental effect of human activity has finally been manifested. As Noss (1995b) pointed out , "many endangered-species conflicts that polarize society ... arguably could have been prevented if management agencies had taken steps to protect adequate amounts and distributions of habitat before populations declined to where listing was legally required." Species at the brink of extinction must be managed intensively, at great cost, and may require immediate and extreme changes in land use that may be strongly opposed by economic interest groups (Noss 1995b). Conversely, a community-level conservation strategy may protect up to 85-90% of species by conserving representative natural communities without a separate inventory of individual species (Noss 1987). Ecosystem-level planning, such as that by the Natural Community Conservation Planning Process of California, may be an efficient way to "avoid the eleventh-hour crises that force choices between losing species and shutting down regional economies" (Mantell 1992, after Noss et al. 1995).

CONCLUSION

Our knowledge of the factors maintaining ecosystem integrity is still incomplete, mainly because of the intrinsic complexity of natural systems. The task of preserving ecosystem integrity is challenging. Even when not influenced by human activities, ecosystems show a high degree of variability, at different temporal and spatial scales, in diversity, structure, and functioning. Such variability reflects changes in the community and physical environment due to internal and external disturbance (Ravera 1991). This inherent variability often makes it extremely difficult to separate the relative effects of natural and anthropogenic perturbations. However, absence of (scientific) evidence should not be interpreted as evidence that environmental impacts are absent. Clear yes-no answers are rarely available and decisions must be made in the face of uncertainty. There are costs in assuming an effect of human activity on ecosystem integrity when there actually is none, but the consequence of assuming no effect when there really is one is often far greater. In a strictly statistical sense, the null hypothesis that human activity does not appreciably affect the integrity of ecosystems can rarely be rejected (Peterman 1990). Waiting for a scientific consensus may delay any decision about conserving the environment, with possibly irreversible consequences. This becomes crucially important if we accept that, in extent and intensity, human exploitation of natural resources seriously threatens the earth's natural capital; too much indulgence in risk-taking could be very dangerous (Woodwell 1989).

Pressure for adequate answers creates a need to devise conceptual tools, such as ecological integrity, to help scientists and resource mangers grasp the complexity of biological systems (Bernstein and Goldfarb 1995). The concept of integrity is far from a panacea for any management problem. Its definition simply reflects the capability of ecosystems, however defined, to support services, including pure aesthetics, that humans value. Ecosystem integrity is not an absolute, monolithic concept, but a multidimensional, scale- dependent abstraction; there is no unequivocal way to apply it in decision making. Measures of integrity must recognize the importance of maintaining processes that support those critical services.

What are the practical implications of these discussions? How should a manager implement notions of ecosystem integrity? The first step is to recognize that this is not the domain of the manager or of the scientist alone. Integrity reflects the ability of ecosystems to sustain services to humans, and the identification of those services can best emerge from multisector partnerships, in which all stakeholders seek agreement on the uses to which an ecosystem will be put, recognizing the linkages with other ecosystems. From such agreement on uses can come the identification of a set of measures that represent the status and trends of those services. A basic research question then arises: how to characterize the relationship between structural features of ecosystems (such as biodiversity or trophic linkages) and measures of functioning? This is an inchoate and nascent area of investigation, but one that holds tremendous potential for advancing the science of management (Daily 1997, Levin 1997, Levin and Ehrlich 1997).

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