This section provides basic guidance for using and conducting economic valuation, including criteria for judging whether valuation is appropriate for supporting decisions. It provides an introduction to the economic techniques used to measure changes in social welfare and describes which methods may be most appropriate for use in valuing particular ecosystem services. Rather than providing comprehensive valuation instructions,it directs readers to additional resources.More generally, it establishes that the valuation of ecosystem services is grounded in a long history of non-market valuation and discusses how ecosystem services valuation can be conducted within established economic theory and techniques.
Economic valuation of ecosystem services can provide decision makers with evidence of the social benefits provided by, and tradeoffs among,regulatory alternatives and other ecosystem management or policy actions.Valuation implies a systematic quantification of benefits and costs realized by society in commensurable (typically monetary) units, using methods grounded in economic theory.1 The strengths of monetary valuation, including aggregation of benefits and comparison of benefits and costs, have led to its widespread use and acceptance by decision makers.
A basis in economic welfare theory is one of the primary distinguishing features of economic valuation. This underlying theory provides a formal structure necessary to link estimated monetary values with changes in net well-being or social welfare. Without this link, there is no guarantee that a monetary measure will have any correspondence to social welfare.
Economic value measures can also be aggregated over affected populations, typically by multiplying the number of affected parties by the magnitude of benefit or harm to each. When monetary measures of ecosystem services have been estimated using methods that share this theoretical grounding, they are internally consistent and directly comparable, regardless of the source or cause of the benefit or cost. This internal consistency and comparability, together with more than five decades of methodological development, are some of the main reasons that monetary values are the most commonly used quantitative measures of ecosystem services value within policy analysis.
The FRMES Guidebook describes a broad suite of approaches for assessing ecosystem services benefits using monetary values and non-monetary benefit indicators. This section focuses on valuation and, reflecting U.S. Office of Management and Budget (OMB) guidance for benefits analysis, defines valuation as economic valuation or monetization.2 Other approaches are referred to as benefit assessments and include the use of benefit indicators that are weighted using risk ranking, multi-attribute utility theory, and multi-criteria decision analysis to reflect public preferences, agency goals, or both and are covered in more detail in the Non-Monetary Valuation Methods section of this guidebook.
When choosing among the benefit assessment techniques presented in this guidebook, consider that economic valuation is rarely the sole input to decision making. In most cases, a variety of other considerations will enter into the decision process that either cannot or should not be monetized. When valuation cannot provide an adequate range of quantitative information required for policy or program analysis, policymakers can supplement or supplant economic valuation with other metrics or approaches. For example, non-monetary benefit metrics or benefit-relevant indicators can be a cost-effective method for comparing alternatives that are difficult to distinguish with monetization.
Agencies can review the Ecosystem Services Assessment Framework Overview to determine which benefit assessment approach best fits their decision-making needs. Choosing a benefit assessment technique(monetary valuation, multi-criteria decision analysis, or quantification of benefit-relevant indicators) requires evaluating the type and accuracy of information needed in a given policy context. For example, legal proceedings (e.g., natural resource damage litigation) may require different standards than collaborative decision making with stakeholders. Some techniques might characterize the effects and tradeoffs of greatest concern better than others. In addition, valuation and non-monetary benefit assessment techniques vary in terms of data and time requirements, so the availability of funding and other resources will influence decisions.
As this guidebook’s ecosystem services assessment framework makes clear, accurate valuation, or any benefit assessment, requires analysts to be able to estimate biophysical cause-and-effect relationships, effects on ecosystem services, and the welfare effects of those biophysical changes. Economic theory imposes specific guidelines on how empirical data can be used to quantify and combine values to prevent the misuse and misinterpretation of data and results.3 Consequently, successful valuation almost always requires the collaboration of ecologists and economists (and possibly others) from the beginning of the process to ensure that the biophysical assessment generates metrics (i.e. benefit relevant indicators) that can be used within economic analysis and to avoid misuse of data when these biophysical changes are translated to measures of economic value. After the metrics have been generated, economists tailor the economic analysis to the ecosystem, social setting, and decision context to ensure that values are estimated from the biophysical changes and that metrics are valid, representative, and useful.
The intent of this section is to provide basic guidance for economic valuation, not to provide comprehensive instructions for all valuation methods. A great deal of research has explained and demonstrated methods and best practices for broadly valuing ecosystem services and non-market resources. This section presents concepts and considerations applicable to choosing an approach to benefits assessment and describes valuation approaches that are most relevant to ecosystem service analysis. Best practices for applications of economic valuation can also be found in the NESP report “Best Practices for Integrating Ecosystem Services into Federal Decision Making.”
Economic values or benefits measure gains in social welfare (or well-being). The economic value of a good or service to an individual or group reflects the increase in well-being that the good or service generates. In the context of ecosystem services, economic values measure gains in social welfare resulting from changes in outputs of natural ecosystems. Because it is impossible to directly observe changes in social welfare, economists measure values by observing the tradeoffs that individuals are (or would be) willing to make. That is, they measure economic value as the amount of one good or service that a person would be willing to exchange for a specific quantity of another good or service, rather than go without the service. When monetary units are used for quantifying a change in welfare, the result is a measure of monetary value. Monetary units are a convenient unit of measure for assessing value, but other metrics can, under certain conditions, be used to represent value.
Valuation often quantifies economic value using individuals’ or groups’ willingness to pay (WTP) for an additional unit of ecosystem goods or services, given the quality of the good or service, the availability of substitutes, and other context variables that affect demand relative to supply. This definition also implies that economic values are based on subjective assessments of individuals regarding their own welfare and the types of tradeoffs that would enhance that welfare. Although the vast majority of economists accept this approach, some competing schools of economic thought propose that individuals may not have or be able to generate preferences that reflect their well-being for all types of goods and services.4 These competing, minority perspectives argue that it may sometimes be appropriate to use group, expert,or political judgment for valuation rather than relying solely on actions or responses of individuals. For example, public spending to enhance wetlands might be used to value a bundle of ecosystem services that the wetland generates—a practice known as public pricing—under the assumption that the government has the information necessary to act in the public interest or the assumption that regulations and government policies reflect public preferences because the public votes for the legislators making the regulations. However, because this is a minority view, valuation methods that rely on such judgments may not be acceptable to all types of decision makers, nor are they approved for use by most government agencies to measure economic benefits.5
Drawing from the above definition of value, WTP measures the amount of money or some other commodity that an individual or group would be willing to give up to obtain a specified quantity of an ecosystem service, compared to a given baseline quantity. Value may also be quantified in terms of willingness to accept (WTA), defined as the minimum amount that a person or group would be willing to be compensated in order to give up a specified quantity of an ecosystem service that they already have or would otherwise get in the future under business as usual. Given these definitions, ecosystem services values are only meaningful for quantified changes in ecosystem services compared to a baseline, and not when summed over entire ecosystems.6
Valuation must account for many variables that can influence the benefits provided by ecosystems, but an issue of particular relevance to ecosystem services is that the value of a change in a system depends on the initial state of the ecosystem or resource. System state (i.e., level of degradation) affects both the magnitude of system response and the significance of that response to users or beneficiaries. For example, a pollution reduction that allows game fish to re-establish in a stream can generate a more substantial increase in recreational fishing opportunities compared to a nutrient reduction in a moderately degraded stream that already contains abundant game fish. In addition, an angler may be willing to pay more per fish to increase her catch from 0 to 1 fish than from 49 to 50 fish, reflecting diminishing marginal returns.
Thus, meaningful economic valuation requires information on the magnitude of the change in ecosystem services from current (and future baseline) conditions and the context-specific value of that change to beneficiaries. The size of the change alone does not dictate value. Rather, the relative scarcity of the good or service (demand relative to supply) determines the value of a change in that good or service. The supply of an ecosystem service is determined by natural ecological processes (often subject to external human influences or stressors). Demand for an ecosystem service is influenced by many factors, including human (subjective) preferences, income, and people’s willingness or ability to substitute other goods and services for the good or service in question.
WTP and WTA are alternative inputs into the calculation of benefits.These distinct ways to measure economic value are appropriate to use in different circumstances because they make different assumptions about who has the right to a good or service.7 WTP and WTA measures may be calculated in a variety of ways, depending on the beneficiary group. For example, one of the most common metrics used to quantify economic benefits or value for individuals is consumer surplus. Consumer surplus is interpreted as the difference between what an individual (or group of individuals) would be willing to pay for a given level of a good or service and what is actually paid, summed over all units of the good or service that are consumed (or used). The change in consumer surplus across all affected individuals may be aggregated to estimate the value of a change in an ecosystem good or service. A parallel measure for producers is producer surplus, which is conceptually similar to economic profit.8
Other metrics that are frequently used as proxies for welfare changes, such as prices, avoided costs, or replacement costs have weaker (or no) ties to social welfare within economic theory. These metrics can still prove useful,because they are relatively easy to measure and can sometimes provide information similar to economic benefit metrics.9 However, the use of these metrics, which do not directly measure economic benefits, must be applied with caution.10
To understand the concerns raised by the use of prices to estimate values, consider a scenario in which a person pays $8 for a bottle of maple syrup. The person loves maple syrup and would have been willing to pay $20 for the bottle. Now suppose someone uses the $8-per-bottle market price and market size to estimate willingness to pay to protect a grove of maple trees.That person will have underestimated willingness to pay to protect the trees for maple syrup production, not to mention all the other values people might have for protecting the trees.
On the other hand, the market value of property saved due to a given flood control investment might be considered acceptable for comparing alternative investments in flood control, as long as decision makers are aware of the pros and cons of this measure. Furthermore, they should be aware of the assumptions that are embedded in use of this measure to reflect social welfare, namely that the value of the property is equivalent to lost welfare. Decision makers should be aware that price and cost data are considered proxies, often for only one service, rather than direct measures of social welfare. As such, they do not provide reasonable or informative metrics of value in all circumstances. An economist trained in valuation can help determine whether and how these proxies may be useful for approximating economic values.
Decision makers often seek the most comprehensive assessment possible in order to promote understanding of how ecosystems provide everything from basic life support to financial and social well-being. Although thoroughness is desirable, complete measurement of all direct and indirect ecosystem services benefits due to a change is rarely possible. Moreover, attempting to be comprehensive increases the risk that benefits will be double counted. The difficulty of putting a dollar value on all ecosystem services is well understood among practitioners and is the reason that OMB guidance suggests that assessments should monetize what is possible to monetize, quantify what cannot be monetized, and describe what can be neither monetized nor quantified for regulatory rule making.11
In an effort to be comprehensive about assessing ecosystem service changes, practitioners have attempted to quantify the total economic value of entire landscapes, biomes, or other very large systems. As noted above, economic value measures are only meaningful for changes in ecosystem services from a known baseline. Moreover, estimated values for small changes cannot generally be scaled up (at least to any significant degree) to calculate values for large changes in systems, such as the systems’ complete loss. Careful selection of ecosystem services for valuation and calculation of values only for changes in ecosystem services rather than entire landscapes can avoid some of these common pitfalls.
Sometimes those seeking to calculate a total value for all or some services provided by a large ecosystem have confused the resulting (generally invalid) estimates with valid measures of total economic value (TEV). This confusion reflects a misunderstanding of the concept. Economists use the concept of TEV to reflect the fact that changes in ecosystem services can simultaneously affect many different types of values, including both use and non-use (also known as passive use) values.12 That is, in the context of ecosystem service valuation, TEV is used by economists to emphasize that a total or complete value measure must incorporate the full range of the types of value and people, including user and nonuser groups, affected by a change. Overlooking one or more of these distinct types of value can lead to large errors and sub-optimal decisions even if many ecosystem services have been valued.The concept of TEV does not imply that one is measuring the total value of an entire landscape or ecosystem. Like all economic values, TEV is only meaningful when well-defined changes from a known baseline are considered.
A typology of ecosystem services values can be used to define TEV (Figure 1). Use values are those that result from direct use (e.g., bird-watching or hunting on site) or indirect use (e.g., flood risk mitigation from proximal wetlands) of ecosystem services or related resources. Non-use values, in contrast, are values that do not require observable use or consumption of the service.13 More broadly, non-use values for ecosystem services are associated with protecting natural assets (e.g., species or ecosystems) because people value the pure existence of these assets, want to pass these assets along to future generations, or think that these assets ought to be protected regardless of human use.14 Option value, which reflects the value of preserving the option to use a resource in the future, had been considered a non-use value. But recent research suggests that option value is more appropriately considered an implicit component of use services, rather than a theoretically distinct and separable component of use value.15
Figure 1. Components of total economic value and relevant valuation methods.
Note: Adapted from R.K.K. Turner, S.G. Georgiou, and B. Fisher, Valuing Ecosystem Services: The Case of Multi-Functional Wetlands (London: Earthscan, 2008).
So, for example, the TEV of a marginal change in waterfowl abundance at a wetland site might include use values realized by waterfowl hunters and birders, along with non-use values realized by those who simply value the existence of these birds. In concept, these and other values are simply summed to calculate the total value generated by the marginal change. Summing must avoid any overlaps in these value estimates that can lead to double counting.
Many economic frameworks are available to evaluate policy choices with and without monetization of ecosystem services values.The most common are cost-benefit analysis (CBA) and cost-effectiveness analysis (CEA).
CBA is the most comprehensive and is specifically designed to quantify total net effects on human welfare. Monetized ecosystem services values are a key component of a CBA, for decisions that affect ecosystems. Although CBA provides more explicit guidance than CEA about whether an action is socially desirable, it also requires that the outcomes expected to generate the major benefits and the costs be monetized.
CEA uses non-monetary metrics of beneficial outcomes as proxies for social welfare (e.g., lives saved, species extinctions prevented) and evaluates the costs of achieving changes in these metrics (e.g., dollars per extinction prevented). CEA is used to identify the options that achieve desired outcomes at the lowest cost, but it does not quantify benefits in monetary terms.
Which approach—CBA or CEA—is most appropriate will depend on the type of decision being made as well as the state of the ecological and social science.
Cost-benefit analysis (CBA) is often required for federal rule making and is the most well-accepted method for communicating the economic desirability and importance of an action to government officials and the public.16 However,it is not without its critics who question its capacity to provide a full accounting of benefits and to help decision makers weigh distributional or equity issues.17
CBA includes a systematic quantification and comparison of well-defined economic benefits and costs resulting from a particular set of actions (e.g., a policy change or management action). It is designed to help society make decisions that increase net economic benefits, considering both present and future outcomes. To that end, CBA requires that key benefits be monetized so that they can be directly compared to costs. For example, CBA may be used to demonstrate that the public welfare is served by a decision (e.g., enacting a regulation or policy) because, when the major benefits and costs are compared, benefits exceed costs.
One of the primary strengths of CBA is that it enables diverse benefits to be evaluated with a common unit (dollars) that directly and explicitly measures social welfare. It is also the only broadly accepted economic approach that is designed to estimate the full range of economic costs and benefits, or effects on human welfare, associated with management or policy actions.
CBA has a long history in economics, and there are now many textbooks and instructional manuals describing its use, along with an established research literature. Individual government agencies have also provided guidance for the use of CBA within policy analysis.18
Cost-effectiveness analysis (CEA) is used to identify the most efficient means of achieving a goal that has been established by legislation, group consensus, technical evaluation, resource management plan, or other means.19 Within CEA, policy benefits are frequently quantified in terms of biophysical indicators that are conceptually linked to human welfare.20 For example, a biophysical change (i.e., a benefit relevant indicator) such as increased groundwater recharge can affect welfare when it reduces pumping costs or enables water use restrictions to be relaxed. However, rather than value groundwater recharge directly, CEA would evaluate the costs of obtaining different possible levels of these indicators (i.e., dollars per million gallons of groundwater recharge).
CEA can be particularly useful for comparing similar alternatives when the outcomes have been linked to a benefit but either cannot be readily monetized or monetization cannot differentiate among sites with variable qualities (i.e., when the effects of an ecosystem services change on human welfare are too subtle or indirect to be measured by valuation studies). CEA is also useful when management or policy outcomes have been predetermined (e.g., to increase groundwater recharge by 20%), and policymakers and resource managers wish to determine the most efficient (or lowest-cost) means of achieving that goal.
Like CBA, CEA has a long history of use throughout federal agencies and other institutions.21 The simplicity of using ecological outcome metrics instead of values can be an attractive option for suggesting benefits, yet it presents unique challenges. For results to be most indicative of social benefits derived from ecosystems, practitioners must develop a process for selecting relevant outcomes and metrics and for ensuring that key tradeoffs of decisions are represented. In addition, if multiple benefit metrics will be aggregated into an index, appropriate methods are needed to maintain the relative social importance of metric changes as they are combined. More broadly, policies or projects chosen on the basis of cost-effectiveness may still generate a net loss of social benefits—nothing in CEA guarantees that benefits will exceed costs.
Another economic tool that is commonly used to evaluate policy effects is economic impact analysis (EIA). This technique is used to estimate the impacts of a given investment (e.g., building a bridge) in terms of changes in economic activity as measured by jobs, economic output, and other metrics. As applied to ecosystem services, EIA is typically used to suggest how investments in ecosystem restoration will generate or maintain economic activity or to show levels of economic dependencies on a natural resource.
Although economic activity outcomes may be desired by policymakers and stakeholders, they are not, strictly speaking, measures of social welfare and are not acceptable for use in CBA or other approaches that evaluate net benefits. To understand why economic activity is not a social welfare metric, consider that spending on any investment comes at the expense of another type investment. Thus, spending will create jobs and economic activity regardless of the investment, although the number of jobs or the level of economic output will vary by the type of investment. Moreover, actions that reduce net social welfare can increase economic activity, particularly in the short run (e.g., cleanup of an oil spill).22
An extensive economic literature addresses theoretical and empirical approaches to valuation of benefits derived from ecosystems.23 The major primary approaches are summarized below using a common typology and selected examples (Table 1).
Table 1. Primary valuation methods applied to ecosystem services.
|Valuation Method||Description||Examples of Ecosystem Services Valued|
|Market Valuationa||Market Analysis and Transactions||Derives value from household’s or firm’s inverse demand function based on observations of use||Fish
Other raw goods
|Production Function||Derives value based on the contribution of an ecosystem to the production of marketed goods||Crop production (contributions from pollination, natural pest control)
Fish production (contributions from wetlands, seagrass, coral)
|Revealed Preferences||Hedonic Price Method||Derives an implicit value for an ecosystem services from market prices of goods||Aesthetics (from air and water quality, natural lands)
Health benefits (from air quality)
|Recreation Demand Methods||Derives an implicit value of an on-site activity based on observed travel behavior||Recreation value (contributions from:
Water quality and quantity
Fish and bird communities
|Defensive and Damage Costs Avoidedb||Damage Costs Avoided||Value is inferred from the direct and indirect expenses incurred as a result of damage to the built environment or to people.||Flood protection (costs of rebuilding homes)
Health and safety benefits (treatment costs)
|Averting Behavior / Defensive Expenditures||Value is inferred from costs and expenditures incurred in mitigating or avoiding damages||Health and safety benefits (e.g., cost of an installed air filtration system suggests a minimum willingness-to-pay to avoid discomfort or illness from polluted air)|
|Replacement / Restoration Cost||Value is inferred from potential expenditures incurred from replacing or restoring an ecosystem services.||Drinking water quality (treatment costs avoided)
|Public Pricing||Public investment serves as a surrogate for market transactions (e.g., government money spent on purchasing easements).||Non-use values (species and ecosystem protection)
|Stated Preference||Contingent Valuation (open-ended and discrete choice)||Creates a hypothetical market by asking survey respondents to state their willingness-to-pay or willingness-to-accept payment for an outcome (open-ended), or by asking them whether they would vote for or choose particular actions or policies with given outcomes and costs (discrete choice).||Non-use values (species and ecosystem protection), Recreation
|Choice Modeling / Experiments||Creates a hypothetical market by asking survey respondents to choose among multi-attribute bundles of goods with associated costs and derives value using statistical models.||Non-use values (species and ecosystem protection), Recreation
Source: Adapted from Table 4.8 in Turner, Georgiou, and Fisher (2008).
a Some typologies consider market valuation a type of revealed preference analysis.
b Most typologies group defensive and damage cost methods under revealed preference techniques. They are separated here because they are more weakly grounded in economic theory than other approaches and can be misused. (See Champ et al. 2003; National Research Council (NRC) 2005; U.S. Environmental Protection Agency (EPA), Science Advisory Board 2009).
Valuation methods fall into two main categories: primary valuation and benefit transfer. Primary methods include market value approaches, non-market revealed preference approaches, and non-market stated preference approaches, including contingent valuation and choice experiments. Defensive and avoided damage costs are discussed as a separate category, even though they include revealed preference approaches, because they are more weakly grounded in economic welfare theory and include techniques that do not necessarily derive values solely from individual preferences, as in the case of public pricing that was discussed above. Benefit transfer approximates economic values for one or more policy sites (where proposed changes will occur) using data or models generated and published for other, similar locations with similar goods (referred to as study sites).
The type of primary valuation that is most applicable and appropriate depends on which component of TEV is being valued (Figure 1).
Further explanation of these methods is provided in Table 1. General summaries of different types of valuation methods and discussions of ecosystem services applications are found in many works.30
Both revealed and stated preference methods have been used extensively since the mid-1980s to estimate values associated with natural resources and ecosystem services. They have been extensively tested and validated and are grounded in large scientific literature.31 However, not all techniques presented here are viewed as equally reliable or appropriate for all decision contexts. In particular, the conditions under which defensive and damage costs provide accurate measures of ecosystem service value are restricted to cases in which substitutability or adaptation is low.32 When defensive and damage costs are appropriate, they are often underestimates of value.
Stated preference methods can be controversial due to their reliance on survey data and hypothetical behavior rather than on observed behavior.33 Because of these concerns, some federal agencies allow the use of stated preference techniques or data only with special permission. Many economists, however, support these methods as a means to quantify values that would otherwise be assumed to be zero or infinite and have countered some criticisms to suggest that CV methods are worth continued use and investigation.34 These methods are considered a valid approach to valuation by many federal agencies and have a long history of use in some agencies such as NOAA National Marine Fisheries Service.35 The well-known NOAA Blue Ribbon Panel also concluded that stated preference methods—when appropriately used—provide information of comparable reliability to many other methods used to support public decisions.36 Balancing these two perspectives, a recent evaluation concludes “that the last 20 years of research have shown that some carefully constructed number based on stated preference analysis is now likely to be more useful than no number [for] cost-benefit analysis.”37 Furthermore, it is almost universally acknowledged that in the absence of these methods, non-use values for ecosystem services cannot be monetized.
Benefit transfer is the use of information from primary studies at one or more study sites to estimate welfare estimates such as WTP or related information at unstudied policy sites. Although the use of primary research to estimate values is generally preferred, the realities of the policy process often dictate that benefit transfer is the only feasible option for estimating ecosystem services values. Benefit transfer is most often used when time, funding, or other constraints prevent primary research and is the most commonly applied valuation technique.38 When used properly, benefit transfer is a useful tool for policy analysis and can be conducted in a manner that minimizes sources of error. But even in the best of circumstances, benefit transfer can be limited by the availability and quality of primary studies on the ecosystem services of interest.39
Benefit transfer involves trading off empirical accuracy for speed and pragmatism (Iovanna and Griffiths 2006). Although benefit transfer enables many ecosystem services benefits to be valued for many sites, the value estimates generated by this method are subject to errors not present in primary value estimates. The size of these errors depends on many factors, including the type of transfer methods applied and the similarity of study and policy sites.40 A significant literature on benefit transfer methods describes methods, limitations, and adaptations.41
Benefit transfer methods are generally grouped into unit value transfers and benefit function transfers. Unit value transfers include the transfer of a single value, for example an average value across multiple studies (e.g., average consumer surplus per bird-watching trip). Benefit function transfers, in contrast, calculate values using an estimated function from empirical research that allows multiple factors (e.g., socio-demographic variables) to be used to adjust the study site value to the policy site. Those applying benefit transfers (and particularly unit value transfers) should be aware of the significant transfer errors that can result, particularly when using unit value transfers across dissimilar biophysical, social, and economic contexts.42 The literature generally finds that more sophisticated benefit function transfers outperform unit value transfers, although unit value transfers can perform satisfactorily if the study and policy contexts (e.g., social factors, geographic and time scales, degree of resource scarcity) are very similar.43
It is standard practice to apply a discount rate to understand the present value of a future stream of ecosystem goods and service benefits. This practice enables current and future values to be compared and aggregated in consistent terms. The discount rate is a function of multiple factors and can be calculated in multiple ways, as described elsewhere.44 As a simplification of this complex topic, one can think of the discount rate as reflecting both the change in the value of money (or a valued commodity) over time and the willingness of current generations to forgo current consumption in order to obtain future consumption.
The choice of discount rate can have a large effect on the outcome of cost-benefit analysis, particularly when evaluating policies with benefits that do not accrue for many years, or that accrue for a very long period of time. The higher the discount rate, the smaller future benefits appear in present value. This issue has become particularly relevant in discussion of appropriate responses to climate change, since the major benefits accrue so far in the future that the present value of those benefits does not compare favorably to the present value of near-term costs, when a typical discount rate is used. The climate change case brings into focus how the choice of the discount rate implies a judgment about the relative importance of benefits received by future generations.
Accounting for intergenerational effects in policy analysis can be difficult and controversial because of competing perspectives on how to set the discount rate. For example, some analysts argue that the discount rate should be low (even zero or negative) in order to promote intergenerational equity. Others argue that using a positive discount rate in policy analysis enables capital investments today that improve the welfare of future generations to a much greater extent than if those investments had not been made.45 The topic remains controversial and as a result analysts often rely on guidance provided by the OMB for regulatory decision making.46 However, that guidance leaves open the possibility that alternative discount rates may be justified under some circumstances. A recent workshop to evaluate the issue found consensus among economists that it may be appropriate to use a declining discount rate for projects with long-term effects, although they did not agree on how to set the discount rate.47
Economic valuation, like any tool, has both appropriate uses and limitations. Valuation is most often criticized for what it omits. However, valuation and CBA are not intended to be the only inputs into decision making and,therefore, they should be judged in terms of whether they can be useful analytic components, even if their results are incomplete.
Among the potential limitations of economic valuation is its anthropocentric (also called utilitarian) perspective, which is often perceived to exclude the intrinsic value of nature.48 More broadly, if an ecosystem service change does not benefit people either directly or indirectly, it has no value within an economic framework.This perspective has led to unease among some, who argue that there is an ethical dimension to ecosystem preservation that is not captured by economic value estimates and that intrinsic values are distinct from other values in that they cannot be traded for other goods and services, as is suggested by a cost-benefit framework. Although an economic framework cannot include all human concerns and, in particular, cannot address the question of whether intrinsic value should be traded off for other uses or values, the TEV framework incorporates such values to the extent that it represents human concerns for protecting nature, independent of human use (i.e., non-use value). That is, because agencies cannot ask nature what it wants but, instead, must ask people to interpret this idea, intrinsic values become existence, bequest, or altruistic values in this framework.
From a practical perspective,many see a lack of diversity in the types of ecosystem services that have been valued. A lack of data or required ecological or other scientific models is often the source of difficulty in estimating defensible economic values.49 Hence, even if one supports economic valuation in principle, empirical applications to ecosystem services can be challenging.
Another concern is that valuation, when used in CBA, can implicitly promote policies that exacerbate social inequities, because when benefits are aggregated, the values of those with higher ability to pay can exceed the values of the poor.50 For example, consider that a choice to build flood protection based on the total value of property protected tends to distribute more resources to wealthy communities than to poor communities. Although ecosystem services value estimates can be extended to quantify effects on different groups or to address equity concerns, most often they do not.
Other concerns can arise from the misuse of valuation tools. Many of the transfers applied in past ecosystem services literature (e.g., particularly in non-economics journals) and in ecosystem services valuation tools have applied methods that would be expected to generate large errors or invalid estimates, particularly due to incorrect aggregation of marginal values, failure to account for spatial connections between ecosystems and their human beneficiaries and their change over time, and other generalization errors.51 For example, benefit transfer studies that purport to value ecosystem services on a biome or worldwide scale are widely considered to be invalid by economists and of little practical value for decision making.52 Similarly, ecosystem services valuation tools that simply multiply a unit value by the area of an ecosystem do not reflect changes in ecosystem services value that will occur with changes in the number of users or beneficiaries, nor do they reflect how values change as resources become scarcer.53 Finally, benefit transfer tools that include values for all ecosystem services may aggregate benefits that should not be aggregated because they compete partially or wholly with one another, meaning that they cannot be simultaneously provided on the same parcel of land (e.g., timber and habitat-related benefits). When tools sum competing services, they double count benefits.
Challenges arising from misuse of tools can be addressed through careful framing of the analysis, and they speak to the need for teaming trained economists with other natural and social scientists throughout all phases of ecosystem services valuation.Many of the shortcomings in prior ecosystem services valuation efforts have been due to the minimal participation of economists in the early stages of the research, when key questions are established, conceptual models are framed, and research methods are determined.
Given its grounding in a long history of nonmarket valuation, ecosystem services valuation should not be viewed as a “new” type of analysis or valuation, but merely as an evolution and reframing of long-established theory and techniques. Ecosystem services valuation provides a systematic means to compare a broad range of social welfare effects using a common (monetary) unit. Furthermore, it applies an internally consistent framework that promotes equivalence among those units. In other words, dollars measure a comparable gain or loss, even if they measure different ecosystem services benefits delivered to different stakeholders. Without this consistent framework, measured units (e.g., dollars) can take on different meanings and fail to represent people’s willingness to trade off goods and services.
Ecosystem services valuation provides information unavailable through other approaches, and is the only means to compare the net social benefits (benefits – costs) associated with policies and actions that affect ecosystem services. At the same time, it is a challenge to design and conduct valuation studies that avoid the common pitfalls of ecosystem services valuation. Consulting an economist with expertise in nonmarket valuation can not only save time and financial resources for the practitioner, but can also ensure that appropriate valuation methods are used and that the resulting data are defensible. When the use of prepackaged (e.g. software) tools is considered, users must understand their limits and verify that the underlying methods conform to economic theory.Although ecosystem services valuation may not represent all philosophical perspectives, well-designed economic valuation studies can support decision making by creating new understanding of the degree to which people benefit from natural systems and by facilitating clear and consistent communication of those benefits.
Arrow, K.J., M.L. Cropper, C. Gollier, B. Groom, G.M. Heal, R.G. Newell, W.D. Nordhaus, R.S. Pindyck, W.A. Pizer, P.R. Portney, T. Sterner, R.S.J. Tol, and M.L. Weitzman. 2013. “How Should Benefits and Costs Be Discounted in an Intergenerational Context? The Views of an Expert Panel.” RFF DP 12-53, Washington, D.C.: Resources for the Future. http://www.rff.org/RFF/Documents/RFF-DP-12-53.pdf.
This paper provides a good reference for the handling of discounting in the case of ecosystem services.
Bateman, I.J., R.T. Carson, B. Day, M. Hanemann, N. Hanley, T. Hett, M. Jones-Lee, G. Loomes, S. Mourato, E. Özdemiroḡlu, D.W. Pearce, R. Sugden, and J. Swanson. 2002. Economic Valuation with Stated Preference Techniques: A Manual. Cheltenham, UK: Edward Elgar Publishing.
This book provides a detailed explanation of the use of stated preference techniques for economic valuation, including the application of these techniques for non-market goods and services such as water or air quality or cultural assets.
Boyle, K.J., N.V. Kuminoff, C.F. Parmeter, and J.C. Pope. 2010. “The Benefit-Transfer Challenges.” Annual Review of Resource Economics 2:161–182. http://www.annualreviews.org/doi/full/10.1146/annurev.resource.012809.103933.
This paper presents a review of benefit transfer literature and a conceptual framework to aid development of unified guidelines for the application of benefit transfer techniques in federal policies.
Champ, P.A., K.J. Boyle, and T.C. Brown. 2003. A Primer on Nonmarket Valuation: The Economics of Non-Market Goods and Resources. New York: Springer.
This book provides clear descriptions of the most commonly used nonmarket valuation techniques and their implementation.
Freeman, A.M., J.A. Herriges, and C.L. Kling. 2014. The Measurement of Environmental and Resource Values: Theory and Methods, Third edition. Washington, D.C.: RFF Press.
This book provides comprehensive coverage of the theory and methods involved in estimating environmental benefits.
Hanley, N., and E. Barbier. 2009. Pricing Nature: Cost-Benefit Analysis and Environmental Policy. Cheltenham, UK: Edward Elgar Publishing.
This book presents a cost-benefit analysis as an economic tool in environmental policy, highlighting special issues posed by environmental management such as valuing ecosystem services.
Holland, D.S., J.N. Sanchirico, R.J. Johnston, and D. Joglekar. 2010. Economic Analysis for Ecosystem-Based Management: Applications to Marine and Coastal Environments. Washington D.C.: RFF Press.
This book discusses the ways that tools of economic analysis inform ecosystem-based management, including applications of ecosystem service valuation.
Johnston, R.J., and R.S. Rosenberger. 2010. “Methods, Trends and Controversies in Contemporary Benefit Transfer.” Journal of Economic Surveys 24:479–510. http://onlinelibrary.wiley.com/doi/10.1111/j.1467-6419.2009.00592.x/full.
This paper synthesizes the literature on benefits-transfer, highlighting methods, trends, and controversies in research and identifying challenges for practitioners.
Kling, C.L., D.J. Phaneuf, and J. Zhao. 2012. “From Exxon to BP: Has Some Number Become Better Than No Number?” The Journal of Economic Perspectives 26:3–26.
This article discusses the controversies surrounding the use of stated preference methods as well as recent advances in stated preference techniques.
McConnell, K.E., and N.E. Bockstael. 2005. “Valuing the Environment as a Factor of Production.” In Handbook of Environmental Economics, edited by K. G. Mler and J. R. Vincent, 621–669. Elsevier. http://www.sciencedirect.com/science/article/pii/S1574009905020140.
This chapter discusses measuring the economic costs and benefits of the environmental changes that influence production.
National Research Council. 2005. Valuing Ecosystem Services: Toward Better Environmental Decision-Making. Washington D.C.: National Academies Press. http://www.nap.edu/catalog.php?record_id=11139.
This report identifies methods for valuing ecosystem services (including non-use values) in the hope of increasing their use in environmental decision-making. It includes case studies and a discussion of uncertainty in valuation.
Ninan, K. N., ed. 2014. Valuing Ecosystem Services: Methodological Issues and Case Studies. Northampton, Massachusetts: Edward Elgar Publishing, Inc.
This book provides examples of different applications of valuation methods for use and non-use values and discusses methods such as benefit-transfer.
U.S. Environmental Protection Agency. 2010. “Guidelines for Preparing Economic Analyses.” EPA 240-R-10-001, U.S. EPA Office of the Administrator, Washington, D.C. http://www.sra.org/sites/default/files/u32/EPA_Guidelines%20_2010.pdf.
This report is one example of agency guidance given for economic analyses.
U.S. Environmental Protection Agency, Science Advisory Board. 2009. “Valuing the Protection of Ecological Systems and Services: A Report of the EPA Science Advisory Board.” Science Advisory Board Committee on Valuing the Protection of Ecological Systems and Services. http://yosemite.epa.gov/sab%5CSABPRODUCT.NSF/F3DB1F5C6EF90EE1852575C500589157/$File/EPA-SAB-09-012-unsigned.pdf.
This report summarizes ecological valuation practices and methodologies, identifies research needs, and recommends next steps for improving the valuation of ecosystem services.
U.S. Office of Management and Budget. 2003. “Regulatory Analysis.” M-03-21, OMB Circular No. A-4. http://www.whitehouse.gov/omb/memoranda_m03-21.
This guidebook details methods commonly applied by agencies in conducting cost-benefit analysis. Much of the guidance is directly applicable to the valuation of ecosystem service changes.