The Meaning of Value and Use of Economic Valuation in the Environmental Policy Decision-Making Process
In developing a perspective and providing expert advice on valuing aquatic and related terrestrial ecosystems, it is necessary to begin with a clear discussion and statement of what it means to value something and of the role of “valuation” in environmental policy decision-making. Environmental issues and ecosystems have been at the core of many recent philosophical discussions regarding value (e.g., Goulder and Kennedy, 1997; Sagoff, 1997; Turner, 1999). Fundamentally, these debates about the value of ecosystems derive from two points of view. One view is that some values of ecosystems and their services are non-anthropocentric—that nonhuman species have moral interests or value in themselves. The other view, which includes the economic approach to valuation, is that all values are anthropocentric.
While acknowledging the potential validity of the first point of view, the committee was charged (see Chapter 1 and Box ES-1) specifically with assessing methods of valuing aquatic and related terrestrial ecosystems using economic methods, an approach that views values as inherently anthropocentric. For that reason, this report focuses on the sources of ecological value that can be captured through economic valuation.1 However, the committee recognizes that all kinds of value may ultimately contribute to decisions regarding ecosystem use, preservation, or restoration. The committee’ s approach is consistent with the approach taken in the international Millennium Ecosystem Assessment,2 which focuses on contributions of ecosystems to human well-being while at the same time recognizing that potential for non-anthropocentric sources of value.
Although this report focuses on the subset of values that can be captured through economic valuation, it is important to emphasize that this subset of values is quite broad; indeed, it is much broader than is often presumed. There are many misconceptions about the term “economic valuation.” For example, many believe that the term refers simply to an assessment of the commercial value of
Unless otherwise noted, use of the terms “value,” “valuing,” or “valuation” in this report refers to economic valuation; more specifically, the economic valuation of ecosystem goods and services.
The Millennium Ecosystem Assessment was launched in June 2001 to help meet the needs of decision-makers and the public for scientific information concerning the consequences of ecosystem change for human well-being and options for responding to such changes (see Chapter 3 and http://www.millenniumassessment.org/en/index.aspx for further information).
something. In fact, the economic view of value actually includes many components that have no commercial or market basis (Freeman, 1993a; Krutilla, 1967), such as the value that individuals place on the beauty of a natural landscape or the existence of a species that has no commercial value. Thus, although economic valuation does not include all sources of value that have been identified or that are potentially important, it encompasses a very broad array of values. In addition, it provides a systematic way in which those values can be factored into environmental policy choices. This chapter provides an overview of economic valuation and the role it can play in improving environmental decision-making. The purpose is first to identify the values that are, and those that are not, captured by the economic approach to valuation and then to discuss how a quantification of these values can contribute to better environmental decision-making.
The chapter is divided into two main sections. The first discusses the role of economic valuation in the policy process and addresses the different meanings and sources of value in this context. The role and importance of quantifying values are discussed next, followed by a discussion of how information about values can be used in policy decisions. Finally, the importance of “framing” the valuation question appropriately is discussed, since the way in which a valuation exercise is defined can have a significant impact on the results that emerge from it.
Given this overview, the following section provides a more detailed examination of economic valuation. The section begins with a description of the “total economic value” framework, from which it is clear that economic valuation includes a wide array of values—many (in some cases most) of which are unrelated to any market or commercial value. This is followed by a discussion of quantifying value using a monetary metric. Two monetary metrics are described, willingness to pay (WTP) and willingness to accept (WTA), and the implications of using one versus the other are discussed. Finally, a discussion of discounting follows because many environmental policy impacts extend over long durations and it is important to incorporate the timing of these impacts into any valuation analysis. Discounting is the approach most commonly used in economic analysis to capture the timing of benefits and costs. The important distinction between discounting as a means of weighing the utility of future generations differently from that of present generations (utility discounting) and discounting as a means of weighing consumption (through benefits and costs) differently at different times (consumption discounting) is highlighted. The chapter closes with a summary of its conclusions and recommendations.
The broad overview of economic valuation provided in this chapter is followed in subsequent chapters by more detailed discussions of the types of ecosystem services that can be valued, the economic methods that can currently be used to quantify those values, and the role of professional judgment and uncertainty in ecosystem valuation.
ROLE OF ECONOMIC VALUATION
Different Sources and Meanings of Value
Given the crucial role that ecosystems and their services play in supporting human, animal, plant, and microbial populations, there is now widespread agreement that ecosystems are “valuable” and that decision-makers ranging from individuals to governments should consider the “value” of these ecosystems and the services they provide to society (Daily, 1997). However, there are different views on what this means and on the sources of that value. The literature on environmental philosophy and ethics distinguishes between (1) instrumental and intrinsic values, (2) anthropocentric and biocentric (or ecocentric) values, and (3) utilitarian and deontological values (Callicott, 2004). In order to place economic valuation in the context of these distinctions, each is discussed briefly below.
The instrumental value of an ecosystem service is a value derived from its role as a means toward an end other than itself. In other words, its value is derived from its usefulness in achieving a goal. In contrast, intrinsic value is the value that exists independently of any such contribution; it reflects the value of something for its own sake. For example, if a fish population provides a source of food for either humans or other species, it has instrumental value. This value stems from its contribution to the goal of sustaining the consuming population. If it continued to have value even if it were no longer “useful” to these populations (e.g., if an alternative, preferred food source were discovered), that remaining value would be its intrinsic value. For example, if the Grand Canyon and the Florida Everglades have intrinsic value, that component of value would be independent of whether humans directly or indirectly use them—either as sites for recreation, study, or even contemplation. Intrinsic value can also stem from heritage or cultural sources, such as the value of culturally important burial grounds. Because intrinsic value is the value of something unrelated to its instrumental use of any kind, it is often termed “noninstrumental” value.
Anthropocentricism assumes that only human beings have intrinsic value and that the value of everything else is instrumental to human goals. To say that all values are anthropocentric, however, assumes that only humans assign value, and thus the value of other organisms stems from their usefulness to humans. Non-anthropocentric or biocentric values assume that certain things have value even if no human being thinks so. Thus, a biocentric approach assigns intrinsic value to all individual organisms, including but not limited to humans. Within this framework, intrinsic value or worth reflects more than humans caring about nonhumans and includes, in addition, the recognition that nonhumans have worth or value that is independent of any human caring or any satisfaction humans might receive from them. For example, a biocentric approach would assign a positive value to an obscure fish population (e.g., the snail darter; see more below) even if no human being feels that it is valuable and thus worth preserving. Clearly, both instrumental value and intrinsic value can be either an-
thropocentric or non-anthropocentric (see Callicott, 2004; Turner, 1999).
Intrinsic value is related but not identical to what economists call “existence value,” which reflects the desire by some individuals to preserve and ensure the continued existence of certain species or environments. Existence value is an anthropocentric and utilitarian concept of value. Utilitarian values stem from the ability to provide “welfare,” broadly defined to reflect the overall well-being of an individual or group of individuals. In this sense, utilitarian values are instrumental in that they are viewed as a means toward the end result of increased human welfare as defined by human preferences, without any judgment about whether those preferences are “good” or “bad.” Existence values still stem from the fact that continued existence generates welfare for those individuals, rather than from the intrinsic value of nonhuman species. As such, there is the potential for substitution or replacement of this source of welfare with an alternative source (i.e., more of something else). In fact, implicit in the economic definition of existence values is the possibility of a welfare-neutral trade-off between continued existence of the species or environment and other things that also provide utility (see more detailed discussion below). Thus, the utilitarian approach implicitly assumes that existence value is an anthropocentric instrumental value that is potentially substitutable.3
In contrast, under the deontological (or duty-generating) approach, intrinsic value implies a set of rights that include a right of existence. Under this approach, something with intrinsic value is irreplaceable, implying that a loss cannot be offset or “compensated” by having more of something else. For example, a human person’s own life is of intrinsic value to that person because it cannot be offset or compensated by that person having more of something else. This approach has its roots in the writings of the philosopher Immanual Kant, who wrote extensively about intrinsic value (e.g., Kant translated in 1987). However, Kant used the concept of rationality to determine the realm of beings that have intrinsic value and rights. He argued that human beings were the only beings who were rational and thus that only human beings have intrinsic value and rights. In this sense, Kant’s views were strictly anthropocentric. Since Kant’s writings, others have suggested alternative criteria for determining the realm for intrinsic value and rights (see footnote 31 in Callicott, 2004) and hence have argued that rights should extend to nonhumans, including animals (either individual animals or species) and in some cases all biological creatures (i.e., all plant and animal life) or the biota collectively. The modern notion of intrinsic value (as used in the context of ecosystem valuation) reflects the notion that rights should be extended beyond human beings (Stone, 1974).4
As discussed in more detail below, the economic approach to valuation is an anthropocentric approach based on utilitarian principles. It includes considera-
tion of all instrumental values, including existence value. Environmental policy and law may also be based on intrinsic value, as exemplified by the Endangered Species Act of 1973. Because it is utilitarian based, economic valuation assumes that the potential for substitutability between the different sources of value that contribute to human welfare. The main categories of value that are not captured by the economic approach are non-anthropocentric values (e.g., biocentric values) and intrinsic values on which the concept of rights is based.
Finally, it is important to keep in mind that economic valuation is based on the notion that the values assigned by an individual reflect that individual’s preferences or marginal willingness to trade one good or service for another, and that societal values are the aggregation of individual values. At any point in time, individual preferences can be influenced by a variety of factors, including culture and information, which can change over time. In addition, an individual’s willingness to trade one good for another will reflect the amount of the goods and services currently available to him, which will in turn depend at least partially on income. If income changes over time, the economic measure of value for an individual can be expected to change as well. For these reasons, the values measured through economic valuation are inherently time- and contextspecific.
Recognition that ecosystems or ecosystem services are valuable, possibly in a variety of ways or for a variety of reasons, does not necessarily imply a quantification of that value (i.e., its valuation).5 In fact, those people who affirm the intrinsic value of ecosystems object to the very idea of trying to quantify the value of environmental goods and services (see, for example, Dreyfus, 1982; MacLean, 1986; Sagoff, 1993, 1994, 1997). For them, that would be as objectionable as quantifying the value of human life. The quantification of the value of ecosystems is by definition anthropocentric since humans are doing it. In addition, it implies a ranking of values (i.e., a statement of which goods or services are “more valuable,” and possibly by how much). Some people object to one or both of these implications of quantification as being analogous to ranking the value of different human beings based, for example, on gender or ethnicity.
However, there are a number of contexts in which quantification of such values may be useful or even necessary, including (1) informing policy decisions in which trade-offs are considered, (2) providing damage estimates for natural resource damage assessment (NRDA) or similar cases, and (3) incorpo-
rating environmental assets and services into national income accounts.6 For example, if an environmental policy decision involves a trade-off in the choice between providing one ecosystem service (such as a particular habitat or an ecological service) and providing another good or service (such as agricultural output), then information about the relative values of these alternative goods or services can lead to better-informed and more defensible choices. This requires a ranking of values, which follows from quantification. A recognition that quantification or valuation may be useful or necessary in informing policy decisions is explicit in the remainder of the committee’s statement of task (see Box ES-1). Given the committee’s charge, the remainder of this report focuses on the role of valuation in the context of policy decisions and improved environmental decision-making. Although not the focus of this study, the committee believes that quantification is also important (in fact, necessary) in the other two contexts as well. In NRDA cases, a quantification of lost value is necessary to determine the compensation that must be paid by responsible parties.7 Similarly, in order to incorporate changes in environmental and other natural assets into national income accounts, these changes must be quantified in a manner comparable to the quantification of other components of national income (Heal and Kriström, 2003; NRC, 1999).
If quantification is deemed to be a useful or necessary input for policy decisions, a particular quantification or valuation approach must be selected. As noted above, given the committee’s charge, this report focuses on the quantification embodied in the economic approach to valuation. In this approach to valuation, the metric that is used to quantify values in nearly all applications is a monetary metric, such as U.S. dollars.8 In the context of ecosystem goods or services that are bought and sold in markets, dollars or some other currency provide a natural metric for quantification since such prices, absent any market distortions, reflect the consumer valuation of that good (see further discussion in Chapter 4). Thus, when policies involve trade-offs between market goods (already valued in dollar terms) and ecosystem services that are not traded in markets, quantifying the value of these nonmarket services using the same metric (e.g., a dollar metric) allows a direct assessment of the trade-offs.
However, the use of a dollar metric for quantifying values is based on the assumption that individuals are willing to trade the good being valued for something else that can also be quantified by the dollar metric. It thus assumes that
Note that the type of quantification that is necessary can vary across these different contexts. For example, NRDA requires a point estimate of the total damages or lost benefits from an environmental reduction in ecosystem services resulting from some event (e.g., an oil spill). In contrast, in a policy context, quantification of the value of a subset of services may be sufficient (see Chapters 5 and 6 for further discussion).
Quantification of values is not necessary if compensation is measured in physical units (e.g., when based on habitat equivalency). However, a habitat equivalency approach to compensation implicitly assumes that the value of the restored or replaced habitat is equivalent to the value of the degraded one.
the good being valued is in principle substitutable or replaceable with other goods or services that are also of value and that money can buy; this reflects the utilitarian principles that underlie economic valuation.9
Role of Valuation in the Policy Process
Although economic valuation requires a quantification of values, the specific design of the valuation exercise should depend on its purpose or the role that it will play in the policy process. One approach is to base policy decisions regarding preservation of environmental resources on moral principles, stemming from a political consensus about what is morally right or wrong. While adherence to moral principles relating to intrinsic value will inevitably involve trade-offs, under this approach these trade-offs are of little or no consequence to the policy choice. If policy choices are to be based on the notion of intrinsic values and rights, then these rights have to be identified, but the values are implied by that identification need not be quantified in order to choose among alternatives (unless the decision to protect one intrinsic value implies a loss of something else with intrinsic value). Thus, with this decision rule, valuation of ecosystem services has no effect on policy choices and hence plays a very limited role (see Goulder and Kennedy, 1997).10
Strict utilitarianism, on the other hand, implies that a decision is based solely on economic efficiency, that is, maximization of the net benefits to society (Goulder and Kennedy, 1997). This decision rule is implemented through the use of benefit-cost analysis (BCA). Economic valuation plays a central role in the application of BCA, since BCA requires an estimate of the benefits and costs of each alternative using a common method (economic valuation) and metric (dollars) so that the two can be compared. The comparison of costs and benefits allows an explicit consideration of the trade-offs that are inevitably involved in most environmental policy decisions. It recognizes that achieving a particular objective or goal such as preservation of a particular ecosystem comes at a cost, since the resources that must be devoted to this preservation are not available for use in providing other goods and services. A typical BCA asks whether the benefits of that preservation are “worth” the costs involved. In this
sense, it ensures that the limited resources used to provide goods and services to society are used in the most efficient way—that is, to achieve the greatest net benefit.
In addition, a benefit-cost approach provides a means of combining heterogeneous views of what is desirable. Although some may prefer preservation of the environment or a particular ecosystem, others may prefer an alternative (e.g., development of the land). These different views can stem from differences in an individual’s net benefits from the alternatives. Those who realize a net gain from preservation would be expected to prefer preservation, whereas those who realize a net gain from the alternative are likely to prefer it. The benefit-cost approach provides a mechanism for combining these disparate views to reach a decision that incorporates both perspectives. Of course, in doing so, it assigns equal weights to the net benefits of all individuals, a property of BCA that may draw criticism (Azar, 1999; Layard, 1999; Potts, 1999).
If BCA is to be used to evaluate environmental policy options, it is imperative that all costs and benefits be considered.11 In particular, for policy decisions that impact ecosystems, the benefits that the ecosystem generates through the various goods and services it provides must be included in calculating the benefits of preserving the ecosystem or the costs (forgone benefits) of allowing it to be degraded. As noted in Chapter 1, failure to assign a dollar value to these benefits (e.g., on the principle that they cannot be valued accurately or that the values are “incalculable”) effectively assigns them a zero value or a zero weight in the calculation of net benefits, implying that changes in those services will not be incorporated into the net benefit calculation (Epstein, 2003).
Political and legal decisions are often made on the basis of information about many sources of value, including intrinsic and moral values, as well as economic values, and some decision rules seek to incorporate different types of values explicitly. For example, decision rules that imply adherence to moral principles or a premise of intrinsic value unless the cost is too high (as under a “safe minimum standard” rule; see Chapter 6 for further information) incorporate concern about both intrinsic value and economic welfare, and implicitly allow some trade-offs between the two. Similar trade-offs are also implied by decision rules that apply a benefit-cost test to environmental policy choices but constrain the decisions to ensure that certain conditions reflecting intrinsic value are not violated. Possible constraints include ensuring (1) that basic notions of justice and fairness are not violated, (2) that populations or levels of critical ecosystem services do not fall below standards necessary to ensure their continuation, and (3) that uncertainties regarding outcomes are not deemed too great. In such cases, information about benefits and costs as determined by economic valuation will be a useful input into the policy decision but will not solely de-
termine it, since the net benefits from the various alternatives will be only one of the factors considered when making a policy choice.
Examples of different weights put on intrinsic values versus utilitarian welfare can be found throughout environmental policies in the United States. For example, the Clean Air Act requires a periodic assessment of the costs and benefits of the act, although it clearly states that the costs or impacts of any standard or regulation promulgated under the act shall not be a basis for changes that preclude the U.S. Environmental Protection Agency (EPA) from carrying out its central mission to “protect human health and welfare.” Thus, information about costs and benefits is intended to inform but not drive policy decisions. In contrast, Executive Order 1229112 required a strict cost-benefit approach to evaluating regulations. The order stated that “regulatory action shall not be undertaken unless the potential benefits to society for the regulation outweigh the potential costs to society.” This order, and a related order (Executive Order 12866), were later replaced by Executive Order 13258, issued in 1996, which replaced the strict benefit-cost criterion for decision-making with a weaker version that instead simply required that the benefits of the regulation justify the costs (OMB, 1996; see also Chapter 4). Under this more recent order, BCA is an input into regulatory decisions but not the sole criterion for them.
Other environmental policies appear to reject more explicitly a consideration of benefits and costs in favor of an approach based on intrinsic value and rights. For example, Callicott (2004) has argued that the protection granted to species under the Endangered Species Act (ESA) is based primarily on principles regarding the duty to preserve species because of their intrinsic value. In Tennessee Valley Authority vs. Hill, the U.S. Supreme Court found that although “the burden on the public through the loss of millions of unrecoverable dollars would [seem to] greatly outweigh the loss of the snail darter…, neither the Endangered Species Act nor Article III of the Constitution provides federal courts with authority to make such fine utilitarian calculations” [emphasis added]. On the contrary, the plain language of the act, buttressed by its legislative history, shows clearly that Congress viewed the value of endangered species as “incalculable” (e.g., Telico dam-snail darter case; U.S. Supreme Court, 1978).13 In response to this finding, Congress immediately amended the ESA to allow at least the possibility of consideration of benefits and costs and to create a committee with authority to grant exceptions to the law’s prohibitions under very limited conditions that consider, but do not simply compare, benefits and costs.
It is clear from the preceding overview that in many policy contexts relating to the use and preservation of environmental resources, some consideration is given to the magnitude of benefits and costs, even though this information is likely to be only one of many possible considerations that influence policy choice. To provide this information, those benefits and costs must be measured, and economic valuation provides a means of measuring them. It is the judgment
of this committee that having the best available and most reliable information about the economic valuation of ecosystem services will lead to improved environmental decision-making. It will allow policymakers to identify and evaluate trade-offs and, if appropriate, incorporate a consideration of these trade-offs into environmental policy design.
Framing the Valuation Question
In order to be useful in the evaluation of environmental policy options, the valuation exercise should be designed or framed to provide the necessary information to policymakers. A number of dimensions are important in framing the analysis. Some of these dimensions are discussed briefly below (see also Chapter 6).
First, it is important to recognize that policy choices, and the benefits and costs associated with them, imply changes in environmental quality or the level of environmental services (e.g., changes in ecosystem goods and services), either positive or negative, and that the valuation exercise is the quantification of the value of those changes.14 Thus, in a policy context, economic valuation is not concerned with quantifying the value of an entire ecosystem (unless the policy under consideration would effectively destroy the entire ecosystem); rather, it is concerned with translating the physical changes in the ecosystem and the resulting change in ecosystem services into a common metric of associated changes in the welfare (utility or “happiness”) of members of the relevant population. Thus, the valuation of ecosystem services should be framed in terms of valuing the changes in those services implied by different policy choices.
A second important dimension of framing is the scope of the analysis. Scope refers to the inclusion or exclusion, by choice or necessity, of certain ecosystem functions or services and/or certain types of value. Thus, a valuation exercise may focus on only a subset of ecosystem services; for example, an exercise might seek to value changes in flood control or water purification services but not changes in the quantity or quality of habitat. Similarly, the valuation exercise may focus (by necessity) on the quantification of certain types or sources of value and may not capture other sources. Although a broader scope provides a more accurate picture of the total impact of the policy change, in some policy contexts a partial approach may be sufficient. For example, if the results of a benefit-cost analysis based on a measure of the partial value of ecosystem preservation imply that the benefits of a particular policy or activity outweigh the costs, then inclusion of additional benefits (by valuing additional services or including additional sources of value) will only reinforce this conclusion (see also footnote 11).
The outcome of the valuation exercise will also depend on its spatial or geographic scale (see Chapters 3 and 5 for further information). Spatial scale has two components. The first is definition of the geographic extent of the relevant ecosystem(s). In defining the physical impacts of a given policy, one can restrict consideration to fairly localized impacts or consider spillover impacts on related ecosystems that are not impacted directly but change indirectly through those linkages.15 Consideration of these indirect impacts will yield a more inclusive analysis, but these indirect effects may be difficult to identify and quantify accurately. In addition, some policies (particularly at the national level) can affect many ecosystems. For example, a categorical exclusion under the National Environmental Policy Act (NEPA) of federal activity in all wetlands 10 acres or less in size will affect the hundreds or thousands of wetlands across the United States. In such cases, the aggregate impact across all affected ecosystems should be valued.
The second component of spatial or geographic scale is definition of the relevant population (i.e., the stakeholders). In estimating the value that individuals place on ecosystem changes, one must identify which individuals (whose values) to include. In other words, what is the relevant population for estimating the benefits and costs of the policy change? For example, in valuing possible damages from a major oil spill, should calculations reflect damages to the local population, to the population within the state, to the population within the nation, or to the world population? Because an oil spill that leads to loss of wildlife may negatively impact those outside the local area who value the existence of the animals, the aggregate measure of damages will generally vary directly with the extent of the population considered (Carson et al., 2001). The appropriate population to include will depend on the perspective of the decision-maker, his or her jurisdiction, and the target population of concern to the decision-maker when assessing the aggregate welfare impacts of the policy change. Thus, local officials may be concerned primarily with the costs and benefits borne by their local constituents, while national policymakers can be expected to take a broader view.
In addition to the spatial or geographical scale, the valuation exercise is also affected by the temporal scale of the analysis (i.e., the period of time over which benefits and costs are distributed). Most policy impacts last for extended periods, and some last (effectively) forever because they lead to irreversible changes. This is particularly likely in the context of ecosystems, where stock effects are important and losses of key ecosystem services may be irreversible. When the benefits and/or costs extend over time, the period of analysis becomes a key factor in determining the results of a valuation exercise. For example, if land conversion for development purposes causes irreversible loss of critical habitat, an analysis that considers only a short time period will not accurately assess the benefits and costs of that conversion. In addition, the analysis should
account for differences in the timing of impacts across alternatives. One approach to this is the use of discounting to weight impacts differently depending on when they occur. The meaning and use of discounting are discussed later in this chapter (see also Chapter 6). At this point, it is sufficient to note that the temporal framing of the valuation exercise—the time period chosen and the method used to reflect differences in the timing of impacts—plays a crucial role in determining its results.
The discussion thus far suggests that the quantification of ecosystem value using the economic approach to valuation can and does play an important role in environmental policy analysis and decision-making. However, the results that emerge from this quantification or the valuation exercise will be influenced significantly by the way in which the valuation question is framed. To provide meaningful input to decision-makers, it is imperative that the valuation exercise seeks to value the changes in ecosystem goods or services attributable to the policy change, that the scope considers all relevant impacts and stakeholders, and that the temporal scale of the analysis is consistent with the scale of the impacts. The results will also depend on a number of methodological and data issues. These issues are discussed in detail in Chapter 4 and illustrated through the case studies provided in Chapter 5.
THE ECONOMIC APPROACH TO VALUATION
Having discussed economic valuation and its role in general terms, a more detailed discussion of the economic approach to valuation follows. As noted earlier, the economic concept of value is based on an anthropocentric, utilitarian approach to defining value based on individual preferences. As such, it does not encompass all possible sources of value. However, it is much broader than the narrow concept of commercial or financial value, and includes all values, tangible as well as intangible, that contribute to human satisfaction or welfare. This broad definition is reflected in the “total economic value” framework that underlies economic valuation and is described below.
The Total Economic Value Framework: Use and Nonuse Values
The total economic value (TEV) framework is based on the presumption that individuals can hold multiple values for ecosystems. It provides a basis for a taxonomy of these various values or benefits. Although any taxonomy of such values is somewhat arbitrary and may differ from one use to another, the TEV framework is necessary to ensure that all components of value are given recognition in empirical analyses and that “double counting” of values does not occur when multiple valuation methods are employed (Bishop et al., 1987; Randall, 1991). It is important to state that the TEV framework does not imply that the
“total value” of an ecosystem should be estimated for each policy of concern. Even a marginal change in ecosystem services can give rise to changes in multiple values that can be held by the same individual, and the TEV framework simply implies that all values that an individual holds for a change should be counted.
In the simplest form, TEV distinguishes between use values and nonuse values. The former refer to those values associated with current or future (potential) use of an environmental resource by an individual, while nonuse values arise from the continued existence of the resource and are unrelated to use. Typically, use values involve some human “interaction” with the resource whereas nonuse values do not. The distinction between use and nonuse values is similar but not identical to the distinction between instrumental and intrinsic value discussed earlier. Clearly, use values are instrumental and utilitarian, but, as noted above, the concept of existence value is not identical to the notion of intrinsic value, because the latter is deontological and includes non-anthropocentric values while the former does not.
Within the TEV framework an individual can hold both use and nonuse values for the services of an aquatic ecosystem. Consider an oil spill on a popular coastal beach resulting in forgone recreational trips to the beach—this is a lost use value. In addition, the oil spill could damage the ecosystem in ways that would not affect beach use and that beach users would never observe. It might, for example, kill marine mammals that live off the beach and are not seen by beach users, and beach users, as well as those who do not visit the beach, might experience a loss because of this ecosystem damage. The loss by those who do not visit the beach would be a loss of nonuse value, though there could also be a loss of nonuse value on the part of beach users. The TEV framework implies that analysts proceed to investigate the potential loss in use and in nonuse values of beach users and in nonuse values of people who do not visit the beach. It is not necessary to estimate the total value of the coastal ecosystem, only the total loss in value associated with the oil spill.
A number of TEV frameworks have been proposed in recent decades (e.g., Bishop et al., 1987; Freeman, 1993a; Randall, 1991). Although varied in detail and application, the distinction between use and nonuse values is a fundamental theme. The TEV framework, as applied to typical aquatic system services for the purposes of this report, is illustrated in Table 2-1. In the discussion below, distinctions are drawn between the components of TEV, but when people hold both use and nonuse values, the literature cited above argues for estimating peoples’ TEV rather than estimating the components and then adding the component estimates to compute a TEV. However, the discussion of valuation methods in Chapter 4 shows that some methods are better able to measure selected components of TEV than others.
TABLE 2-1 Classification and Examples of Total Economic Values for Aquatic Ecosystem Services
Existence and Bequest Values
Commercial and recreational fishing
Scientific and educational opportunities
Nutrient retention and cycling
Shoreline and river bank stabilization
Resources for future generations
Existence of charismatic species
Existence of wild places
SOURCE: Adapted from Barbier (1994) and Barbier et al. (1997).
Use values are generally grouped according to whether they are direct or indirect. The former refers to both consumptive and nonconsumptive uses that involve some form of direct physical interaction with the resources and services of the system. Consumptive uses involve extracting a component of the ecosystem for an anthropocentric purpose such as harvesting fish and wild resources. In contrast, nonconsumptive direct uses involve services provided directly by aquatic ecosystems without extraction, such as use of water for transportation and recreational activities such as swimming. Although nonconsumptive uses do not involve extraction and hence diminution in the quantity of the resource available, they can diminish the quality of aquatic ecosystems through pollution and other external effects.
It is also increasingly recognized that the livelihoods of populations in areas near aquatic ecosystems may be affected by certain key regulatory ecological functions (e.g., storm or flood protection, water purification, habitat functions) (Daily, 1997). The values derived from these services are considered indirect, since they are derived from the support and protection of activities that have directly measurable values (e.g., property and land values, drinking supplies, commercial fishing). For example, mangrove swamps may provide a “storm protection” function in that they may stop coastal storms from wreaking havoc on valuable coastal properties and infrastructure (Janssen and Padilla, 1999). Activities such as reading a book or magazine article about ecosystems, or watching a nature program, are also thought to provide indirect use values.
Many natural environments are thought to have substantial existence values; individuals do not make use of these environments but nevertheless wish to see them preserved “in their own right” (Bishop and Welsh, 1992; Boyle and Bishop, 1987; Freeman, 1993b; Madariaga and McConnell, 1987; Randall, 1991; Smith, 1987). The terms “existence,” “nonuse,” and “passive” use are generally used synonymously in the literature. For the purposes of this report, nonuse values refer to all values people hold that are not associated with the use of an ecosystem good or service. Use values typically arise from a good or service provided by ecosystems that people find desirable. Nonuse values need not arise from a service provided by an aquatic ecosystem; rather, people may benefit from the knowledge that an ecosystem simply exists unfettered by human activity (e.g., Crater Lake). The latter is what was traditionally known as a “pure” existence value in the literature. Other motivations for nonuse values are bequest and cultural or heritage values. The empirical literature generally does not attempt to measure values for individual aspects of nonuse values, but focuses on the estimation of nonuse values irrespective of the underlying motivations people have for holding this value component.
The economic valuation of the impacts of the Exxon Valdez oil spill on the aquatic and related ecosystems of Prince William Sound, Alaska, highlights the importance of nonuse values in natural resource damage assessments and project appraisals (Carson et al., 1992). The Exxon Valdez study revealed that many Americans who have not visited Alaska and never intend to do so nevertheless place high values on maintaining the pristine and unique but fragile coastal and aquatic ecosystems of Alaska. In the context of the Exxon Valdez study, questions were raised about the accuracy with which nonuse values can be estimated (Hausman, 1993; NOAA, 1993). This issue is discussed in greater detail in Chapter 4.
Measurement Using a Monetary Metric: WTP Versus WTA
Economic valuation is concerned with how to estimate the impact of changes in ecosystem services on the welfare of individuals and is based on the principles of utilitarianism. If ecosystem changes result in individuals’ judging that they are worse off, one would like to have some measure of the loss of welfare to these individuals. Alternatively, if the changes make people better off, one would want to estimate the resulting welfare gain.
The basic concept used by economists to measure such welfare gains and losses is rooted in the utilitarian notion that for any individual, the different sources of value that affect the individual’s utility are potentially substitutable; that is, the individual is willing to trade a reduction in one source of value for an increase in another in a manner that leaves his or her overall utility unchanged.
The essence of this approach is to value a change by determining what people would be willing to trade (i.e., to receive or to give up) so they would be equally satisfied or happy with or without the change.
Consider, for example, a case in which a freshwater lake can be restored to enhance sportfishing opportunities. An economic measure of the benefit of such an improvement to recreational anglers is the maximum that anglers would be willing to pay for this improvement in fishing if he or she had to pay. Each angler’s maximum willingness to pay should represent how much money the angler is prepared to give up in exchange for the increase in individual enjoyment gained from the improved recreational fishing. It represents the reduction in income that would be necessary to offset exactly the gain in angler utility resulting from the restoration, thereby leaving anglers at the same utility level as they were prior to any restoration. Maximum willingness to pay could then be aggregated for all anglers who benefit to determine the total benefits of the project.16 This aggregation, in turn, would facilitate an assessment of whether public funds should be spent on the project.
An alternative measure of the value of the improvement in recreational fishing from restoration of the lake is based not on anglers’ willingness to pay for the improvement but rather on the amount they would be willing to accept to forgo the improvement. If the improvement is promised, then failure to provide this improvement (i.e., failure to restore the lake) would reduce the utility of anglers relative to the level they would have attained with the restoration. The value of this loss or the forgone benefit from restoration can be measured by the minimum amount of income that the anglers would be willing to accept as compensation for forgoing that benefit. The increase in income (i.e., the compensation) would have to increase the utility of anglers by exactly the same amount as the reduction in utility stemming from the failure to restore the lake, so that the combined effect would be to leave utility unchanged (i.e., leave the anglers just as well off without the restoration as they would have been with it).
The preceding example illustrates the two alternative measures of value that are used in economic valuation: WTP and WTA. Each measure looks at potential trade-offs between money and the good or service being valued that leave utility unchanged from some base level. They differ, however, in the base level of utility that is maintained when the hypothetical trade-off is made. In valuing an improvement in environmental quality or services, WTP considers trade-offs that would leave utility at the level that existed prior to the improvement (the pre-change utility level), whereas WTA considers the utility level that would exist after the improvement (the post-change utility level).
In some cases such as when valuing small price changes, WTP and WTA measures of value can be expected to be quite close, differing only because of the different income levels implied by paying rather than receiving compensation (Willig, 1976). However, for many environmental goods and services, the
two can be substantially different. In particular, Hanemann (1991) has shown that when valuing changes in the quantities of goods or services available for which there are no close substitutes (including many ecosystem services), the two measures of value can yield quite different results. For environmental improvements, the amount an individual is willing to accept to forgo that improvement will normally be greater than the amount he or she would be willing to pay to ensure it (WTA > WTP).
Because WTP and WTA measures of ecosystem services could differ significantly, a key issue in the use of economic valuation in this context is the choice between these two possible measures of value. As noted above, the conceptual difference lies in the base level of utility that each is designed to ensure. This reflects a difference in the assumption regarding the underlying allocation of property rights or, equivalently, the baseline levels of utility that society collectively agrees to ensure to each individual within that society. Consider again the case of lake restoration. If anglers do not have a right to the improved conditions, then society is not collectively prepared to ensure them a level of utility that includes the restoration. If these anglers want restoration, then in theory they would have to “buy” it from the rest of society. In such a case, WTP is the appropriate economic measure of the value of the improvement. Conversely, if anglers have a right to the improved conditions, then if society wants to use the resources for other purposes, in theory it would have to buy the right to do so from the anglers and pay or otherwise compensate them for failure to restore the lake. In such a case, WTA is the appropriate economic measure of the value of the water quality improvement.
Economic theory, and hence economic valuation, provides no basis for choosing between the alternative property rights regimes and therefore no basis for preferring one measure of value over the other. Property rights are determined collectively by society. In addition, virtually all theories of property rights recognize that they are not absolute or strong but represent only “weak” rights, insofar as they are subject to modification and based on community welfare in ways that strong rights (e.g., a right to life) are not. They are weak rather than strong because they are not considered essential to human dignity in the way that rights to life or to equal protection are (Dworkin, 1977).
Although in theory economic valuation can seek to measure either WTP or WTA depending on the underlying assignment of property rights, it is common to use WTP as an empirically reliable measure. The primary reason is that most of the existing economic methods for estimating values capture WTP but not WTA (see Chapter 4 for further information). The use of WTP may be inappropriate in a given case because of the implicit property rights assumption embedded in it. However, even in cases where WTA would be the appropriate measure, WTP may still be a reasonable proxy for WTA. In theory and practice, the absolute value of willingness to accept usually exceeds the absolute value of willingness to pay (Hanemann, 1991; Horowitz and McConnell, 2002). Thus, WTP can be viewed as a lower-bound for WTA and hence as a lower-bound for the value of the improvement. In some contexts, a lower-bound estimate of val-
ues will be sufficient to inform policy decisions. For example, if the benefits of an increase in ecosystem services exceed the costs when those benefits are measured using WTP, they would also have exceeded costs if measured using a higher WTA. However, if a WTP measure of benefits was lower than cost in a context in which WTA was the correct measure to use, then it is still possible that benefits would have exceeded costs had WTA been used.
In addition to the difference regarding the implicit assumption with respect to underlying property rights, WTP and WTA also differ in another important aspect, namely, the role of income limitations. Clearly, the amount that an individual is willing to pay for an environmental improvement depends on the amount that he or she is able to pay. In other words, WTP is constrained by an individual’s income since he or she could never be willing to pay more than the amount available. WTA, on the other hand, is not income constrained. The amount of compensation that would be required to compensate an individual for accepting a lower level of environmental quality can exceed a person’s income. This difference has important implications in measures of aggregate net benefits. Income constraints imply that, all else being equal, low-income individuals will have a lower WTP than wealthier individuals simply because of their lower ability to pay. This implies that the preferences of wealthy people will get more weight than those of poorer people in net benefit calculations based on WTP. This feature of WTP should be borne in mind when using this measure of value.
Uncertainty and Valuation
Estimates of the values of ecosystem services are frequently somewhat uncertain for a variety of reasons. Chapter 6 explores the major sources and types of uncertainty, indicates which are most significant, and discusses their consequences in ecosystem services valuation. This discussion includes the problems posed by uncertainties about models and parameters, and how analysts and decision-makers can and should respond. Sensitivity analysis and Monte Carlo simulation are discussed as a possible analyst response to model and parameter uncertainties, while risk aversion, quasi option values, adaptive management, safe minimum standards, and the precautionary principle are discussed in the context of use by decision-makers.
Discounting: Utility Versus Consumption
In many ecosystem valuation contexts, the impacts of a particular policy choice will extend over time, and hence an attempt must be made to estimate the costs and benefits not only for current years but well into the future. Deriving an aggregate measure of costs or benefits that reflects their change over time requires an aggregation method that appropriately incorporates the timing of benefits and costs. The most commonly used approach in economic valuation is
discounting, that is, weighting future costs and benefits differently than current costs and benefits when summing over time.
The desirability of discounting future costs and benefits has been the subject of intense debate (Heal, 1998; Portney and Weyant, 1999). The simplest explanation of discounting can be found in the financial context. People generally agree, for example, that accountants are correct to discount future income. If a person will receive an income of $20,000 a year for the next 30 years, most people would agree that it is unreasonable to value that total income at 30 times $20,000. Instead, a more reasonable valuation would be $20,000 for the first year, plus $20,000 discounted by some rate (such as 5 percent) for the second year, plus the amount from the second year, discounted by an additional 5 percent, for the third year, and so on. The rationale for such discounting is the productive power of the economy that converts commodities at one time into a greater quantity of commodities at a later time. If one ignores inflation, then money represents a quantity of purchasing power over economic commodities, and therefore commodities available at an earlier time are worth more than commodities available only at a later time. If the economy remains productive, then (even on a simple level) it is easy to see that money at a later time is worth less than money at the present time because, for example, money this year can be converted into more money in the future by depositing it into a bank to earn interest.
However, the issues raised by the use of discounting in cost-benefit analysis, project evaluation, and ecosystem valuation go far beyond the simple arithmetic of compound interest on bank balances. It is important to realize that two different types of discounting may be practiced—utility discounting and consumption discounting. This distinction is absolutely central, although unfortunately it is not as widely understood. The properties of and justifications for these two rates are quite different, and some of the arguments that apply to one are not relevant in the context of the other (Heal, 2004).
This chapter provides only a brief summary of the underlying issues, which are quite complex and the subject of a massive literature.17 What is normally referred to as “the discount rate” is in fact the utility discount rate, also known as the pure rate of time preference, the social rate of discount, or the social rate of time preference.18 This is the rate to which Frank Ramsey’s famous strictures
apply and indeed those of Roy Harrod as well.19 There is no compelling reason for this discount rate to be positive; the value of the utility discount rate reflects the relative valuations that are placed on present and future generations. If one is convinced that future generations should be valued less than present generations, then a positive utility discount rate should be chosen; otherwise this rate should be zero.
The consumption discount rate is conceptually and operationally different from the utility discount rate. The utility discount rate, as emphasized above, is intended to represent the relative weights put on present and future utilities. It expresses society’s preferences for distribution between generations, with a zero rate representing equal weights for all generations, and a positive rate implying less weight to future people. In contrast, the consumption discount rate represents the weights placed on increments of consumption at different dates. It answers the question, How does one value an extra dollar’s worth of consumption (instead of an extra unit of utility) today relative to an extra dollar’s worth of consumption in the future?
Even if future utilities are valued the same as present utilities (i.e., there is a zero utility discount rate), one may still value an increment of consumption 20 years in the future differently from the same increment today. There are several reasons for this. One reflects changes in wealth or the standard of living over time. Suppose, for example, that people 20 years from now are expected to be wealthier than those today. If the extra utility generated by additional consumption diminishes with income, then providing the additional consumption in the future when people are wealthier will yield less of an increase in utility than providing the same additional consumption today. This suggests that future consumption should be discounted. If this were done, however, it would not reflect a judgment about the relative merits of present and future people, which is what the utility discount rate does. Rather, it would reflect a distributional judgment about the relative merits of extra consumption going to richer or poorer people, quite independent of the dates at which they live. If this approach is accepted, it implies a positive consumption discount rate when living standards are rising over time and, conversely, a negative rate when they are falling.
The distinction between utility and consumption discounting is important in the context of environmental issues (Heal, 2004). One might feel that access to aquatic ecosystem services will decrease over time as a result of human pressures on natural habitat and that, consequently, peoples’ marginal valuations of these services will increase as they become scarcer. As a result, the value of incremental ecosystem services will rise over time and the consumption discount rate to be applied to these will be negative rather than positive. That is to say,
increments in the future will be worth more than those in the present—not because they are in the future but rather because they are being made available at a later date when they are scarcer. This reflects diminishing marginal utility or valuation rather than the result of futurity.
It follows from this discussion that the consumption discount rate is quite flexible and reflects many different characteristics of the underlying problem. If people are concerned with ecosystem goods and services, which are expected to be scarcer in the future than in the present, then the consumption discount rate may be negative, meaning that a unit of consumption in the future would be valued more than a unit at present. If income levels are rising over time, then future income levels will be higher than those at present, so the marginal valuation of income will decrease over time and the consumption discount rate will be positive (i.e., the future should be discounted).
The preceding discussion highlights the existence of two quite distinct concepts of discounting—utility and consumption discounting. It argues that there is no compelling argument for discounting utility, but that there may be reasons for discounting consumption, although the appropriate rate may be positive or negative. When is it appropriate to use the consumption discount rate in ecosystem valuation and when should the utility discount rate be used instead?
In general, the utility discount rate should be used when the policy under consideration is such as to lead to changes in the overall utility or welfare levels of the economy, or at least a significant subsector of it. In economic terms, the utility discount rate is applicable in the context of general equilibrium analyses. The consumption discount rate, on the other hand, is applicable in the context of partial equilibrium problems. These are problems in which only a small part of the economy is being affected by our decisions, and these decisions have only a small impact on overall consumer welfare. Because all of the environmental valuation problems considered in this report are of a partial equilibrium nature, the relevant discount rate to be considered is the consumption rate, which may have either sign. The committee emphasizes that the consumption discount rate is the rate of change of the value placed on an increment of consumption as its date changes. It is not a number that the analyst chooses a priori but one that emerges from the characteristics of the economy, such as whether consumption of the ecosystem good at issue increases or decreases over time. Given this interpretation, one does not argue about whether to discount consumption or at what rate. Discounting consumption—in the very general sense of applying different marginal valuations to increments of consumption at different dates—is unavoidable in the utilitarian framework, and indeed in most other frameworks. One can however argue about the values of parameters that influence, but do not fully determine, the consumption discount rate and in particular determine whether that rate should be positive or negative—that is, whether future costs and benefits should be weighted less or more heavily than current costs and benefits when those costs and benefits are aggregated over time.
SUMMARY: CONCLUSIONS AND RECOMMENDATIONS
This chapter provides an overview of economic valuation and the role it plays in the policy and environmental decision-making process. Although economic valuation does not capture all sources or types of value (e.g., intrinsic values on which the notion of rights is founded), it is much broader than usually presumed. It recognizes that economic value can stem from use of an environmental resource (use values), including both commercial and noncommercial uses, or from its existence even in the absence of use (nonuse values). The broad array of values included under this approach is captured by using the total economic value framework to identify potential sources of economic value. Use of this framework helps to provide a checklist of potential impacts and effects that must be considered in valuing ecosystem services as comprehensively as possible. It reduces the likelihood of omitting key sources of value, as well as the possibility of double counting values. By its nature, economic valuation involves the quantification of values based on a common metric, normally a monetary metric. The use of a dollar metric for quantifying values is based on the assumption that individuals are willing to trade the ecological service being valued for more of other goods and services represented by the metric (more dollars). Use of a monetary metric allows measurement of the costs or benefits associated with changes in ecosystem services.
The role of economic valuation in environmental decision-making depends on the specific criteria used to choose among policy alternatives. If policy choices are based primarily on intrinsic values, there is little need for the quantification of values through economic valuation. In such cases, the “benefit” of preservation is the protection of the right. In such cases, it may still be important to society to know how much protecting that right (e.g., preserving an intrinsically valuable endangered species) would cost—that is, what is being given up to ensure that protection, but there is no need to quantify the benefit of protection. However, if policymakers consider trade-offs and benefits and costs when making policy decisions, quantification of the value of ecosystem services is essential. Failure to include some measure of the value of ecosystem services in benefit-cost calculations will implicitly assign them a value of zero. The committee believes that considering the best available and most reliable information about the benefits of improvements in ecosystem services or the costs of ecosystem degradation will lead to improved environmental decision-making. The committee recognizes, however, that this information is likely to be only one of many possible considerations that influence policy choice.
The benefit and cost estimates that emerge from an economic valuation exercise will be influenced by the way in which the valuation question is framed. In particular, the estimates will depend on the delineation of the changes in ecosystem goods or services to be valued, the scope of the analysis (in terms of both the geographical boundaries and the inclusion of relevant stakeholders), and the temporal scale. In addition, the valuation question can be framed in terms of two alternative measures of value, willingness to pay and willingness to accept
(compensation). These two approaches imply different presumptions about the distribution of property rights and can differ substantially, depending on the availability of substitutes and income limitations. In many contexts, methodological limitations necessitate the use of willingness to pay rather than willingness to accept.
Finally, because ecosystem changes are likely to have long-term impacts, some accounting of the timing of impacts is necessary. This can be done through discounting future costs and benefits. It is essential, however, to recognize that consumption discounting is distinct from the discounting of utility, which reflects the weights put on the well-being of different generations. When the impacts being valued are relatively limited, the discount rate that is used should be the consumption rate rather than the utility rate. The consumption discount rate can be positive or negative, depending on whether consumption is increasing or decreasing. For environmental or ecological services that become scarcer over time, consumption would be decreasing, implying a negative discount rate.
Based on these conclusions, the committee provides the following recommendations:
Policymakers should use economic valuation as a means of evaluating the trade-offs involved in environmental policy choices; that is, an assessment of benefits and costs should be part of the information set available to policymakers in choosing among alternatives.
If the benefits and costs of a policy are evaluated, the benefits and costs associated with changes in ecosystem services should be included along with other impacts to ensure that ecosystem effects are adequately considered in policy evaluation.
Economic valuation of changes in ecosystem services should be based on the comprehensive definition embodied in the total economic value framework; both use and nonuse values should be included.
The valuation exercise should be framed properly. In particular, it should value the changes in ecosystem good or services attributable to a policy change. In addition, the scope should consider all relevant impacts and stakeholders, and the temporal scale of the analysis should be consistent with that of the impacts.
The valuation exercise should indicate clearly whether (1) WTP or WTA measure of value was used, (2) in that context WTP is likely to differ significantly from WTA, (3) in that context WTP is likely to be strongly influenced by income differentials, and (4) use of the alternative value measure instead would likely have led to different policy prescriptions.
In the aggregation of benefits and/or costs over time, the consumption discount rate, reflecting changes in scarcity over time, should be used instead of the utility discount rate.
Azar, C. 1999. Weight factors in cost-benefit analysis of climate change. Environmental and Resource Economics 13(3):249-68.
Barbier, E.B. 1994. Valuing environmental functions: Tropical wetlands. Land Economics 70(2):155-173.
Barbier, E.B., M. Acreman, and D. Knowler. 1997. Economic Valuation of Wetlands: A Guide for Policy Makers and Planners. Geneva: Ramsar Convention Bureau.
Bishop, R.C., and M. P. Welsh. 1992. Existence values in benefit-cost analysis and damage assessment. Land Economics 68(4):405-417.
Bishop, R.C., K. J. Boyle, and M. P. Welsh. 1987. Toward total economic value of Great Lakes fishery resources. Transactions of the American Fisheries Society 116 (3):339-345.
Boyle, K.J., and R.C. Bishop. 1987. Valuing wildlife in benefit-cost analyses: a case study involving endangered species. Water Resources Research 23 (May):943-950.
Callicott, J.B. 2004. Explicit and implicit Values in the ESA. In The Endangered Species Act at Thirty: Retrospect and Prospects, Davies, F., D. Goble, G. Heal, and M. Scott (eds.). Washington, D.C.: Island Press.
Carson, R.T., N.E. Flores, and N.F. Meade. 2001. Contingent valuation: Controversies and evidence. Environmental and Resource Economics 19(2):173-210.
Carson, R.T., R.C. Mitchell, W.M. Hanemann, R.J. Kopp, S. Presser, and P.A. Ruud. 1992. A Contingent Valuation Study of Lost Passive Use Values Resulting from the Exxon Valdez Oil Spill. Report to the Attorney General of the State of Alaska, November. San Diego, Calif.: University of California, San Diego.
Daily, G. C. (ed.). 1997. Nature’s Services: Societal Dependence on Natural Ecosystems. Washington, D.C.: Island Press.
Dreyfus, S.E. 1982. Formal models vs. human situational understanding. Technology and People 1:133-165.
Dworkin, R. 1977. Taking Rights Seriously. Cambridge: Harvard University Press.
Epstein, R.A. 2003. The regrettable necessity of contingent valuation. Journal of Cultural Values 27(3-4):259-274.
Erdheim, E. 1981. The wake of the snail darter: Insuring the effectiveness of Section 7 of the Endangered Species Act. Ecology Law Quarterly 9(4):629-682.
Freeman, A.M., III. 1993a. The Measurement of Environmental and Resource Values: Theory and Methods. Washington, D.C.:Resources for the Future.
Freeman, A.M., III. 1993b. Nonuse values in natural resource damage assessment. In Valuing Natural Assets: The Economics of Natural Resource Damage Assessment, Kopp, R. J., and V. K. Smith (eds.). Washington, D.C.: Resources for the Future.
Goulder, L.H., and D. Kennedy. 1997. Valuing ecosystem services: Philosophical bases and empirical methods. Pp. 23-47 in Nature’s Services: Societal Dependence on Natural Ecosystems, Daly, G.C. (ed.). Washington, D.C.: Island Press.
Hanemann, W.M. 1991. Willingness to pay and willingness to accept—How much can they differ. American Economic Review 81(3):635-647.
Harrod, R.F. 1948. Towards a Dynamic Economy. London: Macmillan.
Hausman, J.A. 1993. Contingent Valuation: A Critical Assessment. Amsterdam: North Holland.
Heal, G.M. 1998. Valuing the Future. New York: Columbia University Press.
Heal, G.M. 2004. Intertemporal Welfare Economics and the Environment. In Handbook of Environmental Economics, Mäler, K-G and J. Vincent (eds.). Amsterdam: North Holland.
Heal, G.M., and B. Kriström. 2003. National Income in Dynamic Economies. Available on-line at http://www.sekon.slu.se/~bkr/nat-inc-april5-02.PDF. Accessed June 3, 2004.
Horowitz, J.K., and K.E. McConnell. 2002. A Review of WTA/WTP studies. Journal of Environmental Economics and Management 44(3):426-47.
Janssen, R., and J.E. Padilla. 1999. Preservation or conversion? Valuation and evaluation of a mangrove forest in the Philippines. Environmental and Resource Economics 14(3):297-331.
Kant, I., (translated by T.K. Abbott). 1987. Fundamental Principles of the Metaphysic of Morals. Buffalo, N.Y.: Prometheus Books.
Krutilla, J.V. 1967. Conservation reconsidered. American Economic Review 57(4):777-786.
Layard, R. 1999. Tackling Inequality. New York: St. Martin’s Press.
MacLean, D. 1986. Social values and the distribution of risk. Pp. 75-93 in Values at Risk, MacLean, D. (ed.). Totowa, Md.: Roman and Allenheld.
Madariaga, B., and K.E. McConnell. 1987. Some issues in measuring existence value. Water Resources Research 23:936-942.
NOAA (National Oceanic and Atmospheric Administration) Panel on Contingent Valuation. 1993. Natural Resource Damage Assessment Under the Oil Pollution Act of 1990. Federal Register 58(10):4601-4614.
NRC (National Research Council). 1999. Nature’s Numbers. Washington, D.C.: National Academy Press.
OMB (U.S. Office of Management and Budget). 1996. Economic Analysis of Federal Regulations Under Executive Order 12866. Available on-line at http://www.whitehouse.gov/omb/inforeg/riaguide.html. Accessed December 2003.
Portney, P.R., and J.P. Weyant. 1999. Discounting and Intergenerational Equity. Washington, D.C.: Resources for the Future.
Potts, D. 1999. Forget the weights, who gets the benefits? How to bring a poverty focus to the economic analysis of projects. Journal of International Development 11 (4):581-95.
Ramsey, F. 1928. A mathematical theory of saving. Economic Journal 38:543-559.
Randall, A. 1991. Total and nonuse values. Chapter 10 in Measuring the Demand for Environmental Quality, Braden, J.B., and C.D. Kolstad (eds.). Amsterdam: North Holland.
Sagoff, M. 1993. Environmental economics: An epitaph. Resources 111:2-7.
Sagoff, M. 1997. Can we put a price on nature’s services? Philosophy and Public Policy 17(3):7-12.
Sagoff, M. 1994. Should preferences count? Land Economics 70(2):127-144.
Sagoff, M. undated. On the value of endangered and other species. Available on-line at http://www.puaf.umd.edu/faculty/papers/sagoff/biodival.pdf. Accessed June 3,2004.
Smith, V.K. 1987. Nonuse values in benefit cost analysis. Southern Economic Journal 54(1):19-26.
Starr, C. 2003. The precautionary principle versus risk analysis. Risk Analysis 23(1):1-3.
Stone, C. 1974. Should trees have standing? Towards a theory of legal rights for natural objects. Los Altos, Calif.: William Kaufman.
Turner, R.K. 1999. The place of economic values in environmental valuation. In Valuing Environmental Preferences, Bateman, I., and K.G. Willis (eds.). London: Oxford University Press.