research on economic valuation

Perspectives on Biodiversity: Valuing Its Role in an Everchanging World (1999)

Chapter: 5: economic methods of valuation, 5 economic methods of valuation.

Chapters 3 and 4 discussed a wide array of services and amenities that biodiversity provides for people who might or might not value its individual components—individual genes, species, and ecosystems—and the diversity of components. Some aspects of biodiversity are valued directly; while others are valued for their contributions to ecosystem support and, hence, to sustainable production of things that are valued directly. The economic value of biodiversity has its place in the policy-making process. Although biodiversity might well have substantial economic value, compared with alternative consumptive resource uses, economic value does not tell us everything we need to know about the value of biodiversity.

Economic valuation is an attempt to provide an empirical account of the value of services and amenities or of the benefits and costs of proposed actions (projects or policies) that would modify the flow of services and amenities. Economic valuation provides a utilitarian account, that is, an account of contribution to the satisfaction of human preferences (see chapter 4 for a detailed discussion). Therefore, it provides a particular perspective on value—in this case, on the value of biodiversity. Utilitarians might object to some aspects of the economists' utilitarian account: to produce an economic account of contribution to preference satisfaction, a particular kind of structure has to be introduced into the analysis, and utilitarians will not always endorse the process or the results. In addition, there are many nonutilitarian perspectives on value (see chapter 4 ), which deserve consideration on their own merits.

Theoretical Foundations

Welfare-change measurement.

The foundation of benefit-cost analysis (BCA) is welfare-change measurement: the benefit from some proposed action is the money-related welfare change that it generates. The concept of benefit is an increase in welfare, that is, preference satisfaction; and welfare change is measured in terms of money. Valid money measures of welfare change can be defined conceptually and can be estimated with reasonable accuracy, precision, and reliability; and individual welfare changes to arrive at social benefits and costs can be added up. Skepticism about any of those claims, in general or in the specific application to biodiversity, suggests that caveats should be applied to the interpretation of benefit-cost information or its use in policy decisions.

The conceptually valid measures of welfare change are willingness to pay (WTP) for benefits and willingness to accept (WTA) for costs. WTP is the amount of money that someone would willingly pay to get a desired good, service, or state of the world rather than go without; WTA is the amount of money that would induce someone to willingly give up the good, service, or state of the world. Those measures are readily defined in market terms—WTP is the buyer's best offer, and WTA is the seller's reservation price (the price at which the seller will hold rather than sell)—but they are by no means restricted to commodity markets. Some people are willing to pay substantial amounts of money for improvements in the quality of their life. Some would willingly accept a lower level of amenities if compensated with money; for example, some would willingly move to an undesirable location if promised a large enough pay raise.

For BCA of a policy proposal, aggregate benefits are defined as the sum of WTP figures for all those who stand to gain from the proposal. Aggregate costs are the sum of WTA figures for all who would provide goods and services or bear disamenities if the policy proceeds. Some critics object to aggregating benefits or costs that accrue to individuals, on the grounds that individuals with greater income and wealth tend to have greater WTP (or WTA) and that simple aggregation makes no attempt to correct for this or to place extra weight on things that benefit the disadvantaged.

Given that many proposals promise benefits and costs continuing well into the future, the "bottom line" of the BCA is expressed as net present value, that is, the difference between the sum of present and future benefits and the sum of present and future costs, all discounted to the present. The practice of discounting has been controversial in some circles, especially in the context of environmental projects and policies (for example, Daly and Cobb 1989), where it is claimed that discounting tends to trivialize the demands of future generations for present conservation (see box 5-1 ) . That argument has been winning fewer converts in recent years (Heywood 1995), as economists have been reminding us

that realistically high discount rates discourage wasteful investments that would actually harm future prospects. Current writers are skeptical about the wisdom of using low discount rates to achieve policy goals, preferring more direct approaches to the concerns of environmentalists. For example, Howarth and Norgaard (1991) argue that balancing equity among generations should be addressed by intergenerational transfers of resources, and Farmer and Randall (1997) suggest that targeted conservation policies provide the appropriate remedy to the extent that particular natural resources are both necessary for human welfare and threatened with exhaustion.

Cost-effectiveness analysis can be useful in guiding decisions toward the

most efficient way of meeting specified goals. However, it does not provide estimates of values. If there are several ways of accomplishing a particular and well-specified goal, cost-effectiveness analysis compares the costs of the various approaches; the most cost-effective is the one that accomplishes the goal at the lowest cost. If different approaches would achieve different quantitative levels of performance, cost-effectiveness might be expressed as cost per unit of performance (for example, cost per acre preserved or cost per nesting pair saved). If the policy-maker is confident that the different approaches are otherwise equivalent in terms of the results achieved, choosing the most cost-effective approach is justified.

Categories of Value

WTP and WTA for some natural resource or amenity are equivalent to its total economic value. However, humans use and enjoy natural resources and amenities (as they do other goods and services) in a variety of ways. At one extreme, natural resources can provide commodities that are purchased and consumed directly; at the other extreme, people might enjoy satisfaction that a particular habitat is being maintained at high quality. Both kinds of use generate economic value, but it is likely to be expressed in different ways and via different institutions for commodities, in terms of quantities taken and prices paid and for habitat quality, perhaps via voluntary contributions to conservation organizations.

Total economic values include all the several kinds of economic values that have been identified by economists. Total economic value is the WTP for a change in the state of the world. To impose some order and consistency, the following relatively simple classification of economic value is gaining ground among economists.

Use value is generated when a person uses an environmental service actively, typically by consuming it directly or combining it with other goods and services and the person's own time to "produce" an activity that generates utility. Recreation experiences, for example, are produced by combining on-site amenities with travel services, recreation equipment, and the participant's time. Use value is likely to be reflected (at least in part) in behavior such as purchases and visits.

Use value, naturally, includes the expected value of future use. If uncertainty attends future availability of an amenity or future demand for it and potential users are risk-averse, use value under uncertainty can include option value , the value of assurance that things (such as biodiversity) that are available now will still be available when we need them, and quasi-option value , the value of waiting to decide on the disposition of an asset (such as whether to build houses on Camp Pendleton—see the case study in chapter 1 ) motivated by the possibility that we will be able to make a "better" decision later, perhaps because we will have more information. When institutions provide opportunities for individuals to secure options for future use, these kinds of value might be reflected in behavior.

Passive use value captures the idea that people might enjoy satisfaction from "just knowing" (that is, enjoying the assurance) that a particular habitat is being maintained in good condition. There is no general expectation that passive-use value involves overt activities or is reflected in behavior. However, contributions to voluntary organizations that provide habitat preservation and political support for pro-habitat policies are consistent with passive-use value.

Together, use value, option value, and quasi-option value make up total economic value. It also includes bequest value, in that bequest motives assume that one's heirs will enjoy use or passive use. Total economic value includes all the kinds of economic value. There is no claim that economic value, however, constitutes the totality of value. As chapter 4 has made clear, there are many ways of valuing , but, total economic value then represents a comprehensive application of the economic way of valuing.

Methods of Valuation

Valuation relies on detailed information from the natural sciences . We might value an environment as an asset, in which case its value would be the net present value of the services that it provides and will provide. Alternatively, we might evaluate some proposed action (a project or policy); value would then be the net present value of the change in services that the environment will provide minus the cost of implementing the proposed action. Either way, valuation requires detailed knowledge of the service flows of the environment, of the costs incurred in preparing these services for human enjoyment, and of the responsiveness of service flows and costs to human interventions (Randall 1987 and NRC 1997 provide conceptual models of the valuation process). Much of that information must originate with experts whose specialties are far from economics, for example, ecologists and hydrologists. Economic valuation depends heavily on information that is fundamentally noneconomic.

Valuation also requires evidence of WTP and WTA . Evidence of WTP and WTA varies along two dimensions of quality: consistency with the conceptual framework of welfare-change measurement and reliability of the data themselves. For example, data generated by market transactions are convincing in at least one respect—paying money is the sincerest expression of WTP and accepting money and relinquishing an amenity constitute the sincerest expression of WTA. But the data might, for a variety of reasons, fail to measure the correct value concept. Price typically indicates marginal value (literally, the value of the next unit more or less than the status quo quantity—a small change); but a proposal might involve nonmarginal (big) changes. In addition, market distortions of various kinds might distort prices, markets might be incomplete or otherwise imperfect, and the environmental service involved might be nonmarketed. Data generated by contingent valuation or contingent policy referendums often can (because a researcher controls the valuation context) be addressed to the right value mea-

sure, but still this might raise doubts as to whether contingent payments and votes are reliable predictors of behavior. Valuation researchers are often faced with one or another form of this dilemma: "harder" data might depart from the ideal, and conceptually valid measures might be "softer". In some cases where hard data depart from the ideal value concepts, economists have developed ingenious methods of inferring the ideal values; however, there is always a risk that they will be forced to substitute assumptions for evidence and structure for information. The resulting value estimates will be to some degree artifacts of the methods used and the research decisions made.

Direct and Indirect Evidence from Markets

It is hard to imagine a market for biodiversity as a whole, but its various components are routinely marketed. Consider a biodiverse forest. Timber, fuel wood, and some nonwood products can be produced and sold. The forest can provide catchment for water that is valued by downstream farmers and urban residents. The forest ecosystem can harbor genetic resources with commercial potential, for example, rare species that might be of pharmacological interest or wild species that are precursors of modern, commercially important plant varieties. Recreationists and nature-lovers can devote resources (money and time) to visiting the forest. People can buy homes near the forest to have access to its amenities. The productivity and value of these various activities depend on how the forest is managed, so proposals that affect forest planning and management will generate costs and benefits that are reflected, to various degrees, directly or indirectly in markets. The Pacific Northwest case study in this chapter provides a detailed example.

Market Demands and Prices

For commodities that can be sold in quantities that are small relative to the total market, the economic value that can be assigned to a decision to sell or preserve is simply the product of the market price of the commodity and the quantity. For example, if 20 acres of old-growth timber is reserved from the market to protect a pair of spotted owls, one estimate of the cost of this decision is the product of the volume of the timber and the market price per unit volume. That holds as long as the quantity is so small that its removal from the market does not affect the market price of timber generally.

But consider the Pacific Northwest (see case study, this chapter), in which the area of federal forest taken off the market to protect the spotted owl and other threatened species had accounted for some 10% of the nation's softwood lumber production. The quantity of timber removed from the market was clearly large enough to affect the market price for timber in much of the country. For such a decision, an estimate of the economic cost must consider not only the change in

the quantity of timber marketed from that area, but also the change in the market price per unit volume of the remaining marketable timber (from before the decision to after the decision). Such an estimate must also consider what economists refer to as "substitution effects". The changes in quantities and prices that result from a decision of this magnitude affect the market price of timber in other areas. Timber producers in forests other than those immediately affected by the decision (for example, the southern United States) respond to the change in timber price by changing the quantities of timber that they put on the market, thus causing further changes in the price of timber. Additional complications include the effects of the changes in timber price on the marketing of such substitute products as steel and plastics and the modifying effects of time as these various factors work through the marketplace. In sum, calculating the effects of decisions that affect market prices is not easy, but it is conceptually feasible.

The goal of valuation—measuring net present value—introduces two complications. First, "net value" requires that any costs associated with using the resources, such as, timber-harvesting costs, be subtracted from gross value. Second, "present value" requires prediction of future demands for environmental services. In the cases of timber, water supply, and genetic materials, the forest augments the supply of things that are valued as factors in production. So the demand for forest products is a derived demand, which complicates predictions of demand: the analyst needs to be concerned with demands for the final products (houses, irrigated crops, and pharmaceutical products) and with the supply of other things that might serve as substitute factors in their production.

The idea of substitutes suggests another approach to valuation: when it is hard to observe market demands for forest products directly, the analyst might look to market evidence concerning substitutes. For example, the avoidance cost method might value improved water quality by observing the household water-filtration costs avoided, and the replacement-cost method might value increased water catchment by calculating the cost of additional reservoir capacity that would serve the same purpose. In both cases, the methods provide an upper-bound value for the particular services they address: the services cannot be valued at more than the cost of avoiding the need for replacing the service with a perfect substitute, but they could be valued at less than that, in the event that effective demand would not clear the market for these services at these prices. The Quabbin Reservoir and Lake Washington case studies in chapter 6 illustrate this.

Travel-Cost Methods

Recreationists spending their money and time to visit the forest leave a trail of indirect evidence about their WTP for the services and amenities that it provides, and travel-cost methods attempt to tease out this WTP. The weak-complementarity assumption, of course, limits the travel-cost method to estimating the

use values associated with site amenities. The simplest travel-cost models posit simply that the number of visits, at a given level of site quality, is a function of travel costs and socioeconomic variables, where travel cost is a proxy for the ''price'' of visits and includes costs of distance traveled and time spent in traveling. Substitute sites and activities typically are included in arbitrary fashion or assumed to be of little import (formally, this is accomplished via assumptions of separability in the utility function). A large literature attests to the difficulty that researchers have experienced in estimating the cost of travel time (Bockstael 1995), but this is symptomatic of a general difficulty: it is inherently difficult for researchers to observe the cost of a visit, that is, the value opportunities foregone to make the visit (Randall 1994). If one assumes a relationship between the quality of on-site amenities and the costs of goods and services used in traveling to the site, the value of an increment or decrement in site quality is measured as the integral between demands for visits at the with-proposal and without-proposal levels of site quality.

The random-utility model (RUM) has become the travel-cost model of choice (Bockstael 1995) because its systematic treatment of substitute sites allows it to characterize site quality more completely. RUM models are therefore more useful than basic travel-cost models for valuing changes in levels of environmental amenities. Their disadvantage arises from their substantial information needs, which in practice often lead to the use of very large data sets and simplifying analytical assumptions that impose rigidities; thus, estimates based on travel-cost models are to some degree influenced by researchers' analytical choices.

When travel-cost models are used to predict number of visits, validation is relatively simple, and several well-known models have performed well (for example, Bockstael and others 1987). However, direct validation of the value estimates obtained with travel-cost models is impossible; the best one can do is test for convergence of the results of travel-cost methods and the results of alternative approaches, such as contingent valuation, and such tests have provided some empirical evidence of convergent validity.

Hedonic Price Analysis

Hedonic price analysis separates the factors that contribute to prices to identify the contribution of those based on environmental amenities. Imagine a good with several important or desirable features, such as a house or automobile. It is a reasonable hypothesis that the price of a particular house or car reflects its particular characteristics. If a statistical analysis succeeds in explaining the price of a house as a function of its characteristics and one of those characteristics is the level of environmental amenities, then the marginal (small) impact of a change in an amenity level (a trait that makes it attractive) on the house price should provide evidence of this amenity value. This is the intuition behind hedonic price analysis. A hedonic price function, relating house prices to characteristics, is

estimated. Typically, three kinds of characteristics are used: on-site characteristics, such as the number of bedrooms; neighborhood characteristics, such as school quality; and environmental amenities, such as access to a biodiverse forest. The first derivative of the hedonic price function with respect to the environmental characteristic of interest is its hedonic price (or marginal implicit price), a measure of the marginal value of the amenity.

The literature suggests that hedonic price analysis has succeeded, in a fairly wide range of circumstances, in generating plausible estimates of marginal hedonic prices for various housing characteristics, including environmental amenities. To value nonmarginal changes in amenity levels, however, it is necessary to estimate hedonic demands, that is, demands for amenities. The literature reports many attempts to find conceptually valid methods of identifying hedonic demands, but no method has proved generally acceptable. Hedonic price analysis is often effective for valuing marginal changes in the levels of environmental amenities that can be accessed via, say, choice of home site but cannot generally be used for valuing nonmarginal environmental changes. The assumptions underlying the method limit its application to a subset of use values; for example, a housing-price hedonic analysis will measure use values associated with home site amenities, but not values that can be accessed regardless of exactly where one lives.

Evidence from Self-Reports

If we design and ask the questions with enough care, perhaps people can provide reliable evidence of amenity values by telling us their WTP or WTA directly or by telling us what they would do (for example, buy or not buy or vote yes or no) if given well-specified choice situations that we construct to generate data that we can analyze to infer their WTP or WTA. That is the intuition behind contingent valuation and contingent-choice experiments. The great advantage is that the researcher controls the context of choice, which makes it possible to estimate total economic value, passive-use value, and various use values that can elude the methods that use market-generated evidence, directly or indirectly. A further advantage is that information can be obtained to value amenity levels beyond the existing range; if it can be described by the researcher and comprehended by the respondent, it can be valued. The potential disadvantages lie in the self-reported nature of the data: some people might seek to answer strategically, some might answer carelessly, and some might struggle mightily (but hopelessly in the end) to provide valid responses to questions that cannot be answered meaningfully. Economists, who are weaned on the admonition to "watch what people do, not what they say", approach these methods with a well-developed skepticism; yet the results, although mixed, have been encouraging enough to stimulate a proliferation of applications.

The techniques require primary data collection in a survey or experimental

context. With rapid advances in information and communication technologies and increasing synergism among research programs in, for example, economics, social psychology, and marketing, it is reasonable to expect vigorous innovation in research design and data collection methods. In this report, we use the standard categories of contingent valuation (in which responses to one or a few choice questions provide the basic data for valuation) and contingent-choice experiments (in which value is inferred from responses to a sometimes long sequence of pairwise choices). The basic project underlying the methods is to learn about value from people's self-reports; and as development and testing of these methods proceed, we can expect new approaches to emerge and existing categorizations to become obsolete.

Contingent Valuation

The essential elements of a contingent-valuation (CV) exercise are a description of the default and alternative situations (respectively, what you get if the proposal fails and if it passes), the institutional environment, the valuation question, and the policy-decision rule: How does the answer to the valuation question affect whether the proposal passes or fails? (See the Grand Canyon flush case study below.) The valuation question can be continuous (or open-ended); for example, What would you be willing to pay? Or it can be in the form of a dichotomous choice; for example, Given the stated cost to you and the policy-decision rule, would you vote yes or no? (Alternatives in common use are, Would you buy it or not?; Would you donate to the trust fund or not?) The different forms of the valuation question require different analyses to estimate WTP or WTA; for example, the results of the dichotomous form are usually analyzed with some kind of RUM (Hanemann 1984). With different policy-decision rules, they imply different kinds of incentives for truthful responses (Hoehn and Randall 1989).

There is already an extensive literature of CV applications, and attempts to validate CV include tests for internal consistency and tests of convergence with value estimates obtained with different methods. Encouraging results have been obtained (for example, Carson and others 1996; Smith and Osborne 1996), but critics have raised enough doubts (for example, Hausman 1993) for CV to remain controversial. A 1993 report by a prestigious panel (Arrow and others 1993) failed to settle the issues when it endorsed CV in principle, even for measuring passive-use values in environmental-damage litigation, but announced a long and demanding list of standards that a valid CV should satisfy. CV that would meet the panel's standards would be prohibitively expensive in most applications, and, as methodological innovation and the accumulation of evidence proceed, the process of rethinking the panel's recommendations is beginning.

One of the panel's recommendations deserves highlighting here. In keeping with a good deal of professional opinion, the panel concluded that CV could not

be endorsed for estimating WTA directly—that whereas WTA is the appropriate measure of value for decrements in environmental services, considerations of reliability lead to the recommendation that, instead, self-reports of WTP to avoid loss can substantially understate WTA (Hanemann 1984). The panel's recommendation would have the effect, therefore, of undervaluing the losses from destruction of unique ecosystems.

Contingent Choice Experiments

Open-ended CV sets a rather difficult task for respondents (announcing a dollar value of some nonmarketed amenity), and dichotomous-choice CV sets a simpler task (announcing whether a proposal is accepted or rejected at a specified cost) but collects only one or two valuation data points. It might be argued that progress could be made by having respondents make a larger number of simple pairwise choices. That is the motivation for contingent-choice experiments, in which data generated by a sequence of pairwise choices are analyzed with RUMs to generate value estimates. As Adamowicz and others (1994) demonstrate, these methods have another potential advantage: contingent-choice and actual-choice data can be combined to extend the range of data points and to test for consistency between the two kinds of data. The methods also have disadvantages: a long sequence of pairwise choices can tax respondent's patience, and the RUM analytical framework imposes rigidities on the analysis.

Contingent-choice experiments are a fairly recent development, so the evidence on their performance is rather thin. Initial applications have emphasized amenity-use values, but there is no inherent reason why they could not be used to estimate passive use and total values, and (given the CV controversies) current research in environmental-damage assessment is heading in that direction.

Case Study: The Grand Canyon Flush

Dams and reservoirs on major streams affect downstream conditions by changing water flows, water temperatures, sediment loads, and the character of stream-bottoms and beaches. A large-scale test of the potential for reestablishing stream-bottom and beach characteristics in the Grand Canyon of the Colorado River was conducted in the spring of 1996. A week-long surge of water through the canyon was provided by opening the gates on the Glen Canyon Dam just upstream of the Grand Canyon. Initial results of the experiment were evident almost immediately, even while the "grand flush" was in progress.

A major purpose of the experiment was to determine whether sandbars along the river could be restored for recruitment of riparian trees and shrubs, an important element in the canyon's ecosystem; a finding before the experiment indicated that the dam-caused lack of springtime floods had reduced the energy in the river needed to lift bottom sediments onto adjacent sandbars. An additional goal was

to improve habitat for some native species of fishes, at least one of which is endangered. To meet those goals, water storage behind the dam had to be reduced, thereby reducing potential water supplies to downstream users and power generation. The interests affected by the flush included water and power users; recreationists, especially rafters who use the sandbars; sport fishers who value the cold-water trout fishery below the dam; Indians living along the Grand Canyon; and species whose habitat is affected by river flows and the character of the canyon's ecosystem.

The gates on the Glen Canyon Dam in northern Arizona were opened on March 26, 1996, after over 10 years of study. The dam was built primarily to control water flows to meet domestic, municipal, and irrigation needs in the downstream states of California, Arizona, New Mexico, and Nevada and the upstream states of Colorado, Utah, and Wyoming. But the dam was also an important producer of hydroelectric power. After construction of the dam and filling of the reservoir, other concerns grew in importance, including river rafting and protection of species, some of which were endangered fishes. The effects of the dam on the character of the river—the smoothing of normal seasonal variations in flows through the Grand Canyon—were also recognized. Proposals were made to open the dam's gates occasionally to allow flood-like surges of water to rebuild the sandbars along the river banks. But lowering water levels in the reservoir was also seen as a threat to meeting the needs of water users and as expensive in terms of reduced power supplies.

Lake Powell (behind the Glen Canyon Dam), Lake Mead, and a number of smaller reservoirs on the Colorado River and its tributaries, hold about 4 times the average annual flow of the Colorado River. This storage is used to even out the year-to-year and seasonal supply of water to various users, divided between the upper and lower basin states by interstate compact and court decrees. In view of the history of large variations in runoff and of periodic droughts, the Glen Canyon Dam was built upriver of Grand Canyon National Park in part to provide water to the upper basin states and in part to regulate flows into Lake Mead, which supplies much of the water to users in California.

The dam is also a major source of power, although generating income from this power has always been secondary to supplying water. Congress specified that, although power production was not to interfere with supplying water, maximal power at firm rates was an objective. Without the power production, the dam would have been seen as uneconomical and might never have been built (NRC 1996:11–3). Once the dam began to operate, the smoothing of seasonal variations in water flows had major effects on the river as it flowed through Grand Canyon National Park. The dam stored and substantially reduced the sediments carried by the river. In addition, the elimination of high flows in the spring reduced the

ability of the river to move coarse sediments that reached the river from tributaries below the dam. The result was a loss or reduction in height of the many sandbars along the river that provided camping sites for recreationists and backwater habitat for native fish. The release of water from the pool behind the dam also lowered the average temperature of the river, especially in the stretches just downstream of the dam. This favored introduced species of trout but militated against native fishes, such as the humpback chub.

Use of the river was also changing, especially with the rapid increase in recreational river rafting. The increase in average daily flows in the summer increased the opportunities for rafting, and maintaining relatively steady flows, as opposed to the fluctuations that come with providing power on demand, is generally seen as improving the experience of white water rafting. The loss of sandbars on which to camp is seen by rafters as a negative factor. The river downstream of the dam also became a valued cold-water fishery for trout. Those factors assumed greater importance relative to supplying electric power in the early 1980s. The Bureau of Reclamation (BOR) and the Western Area Power Administration (WAPA), the federal agencies responsible for operation of Glen Canyon Dam and marketing power from it, however, still saw their responsibilities as meeting the goals of supplying water to users and electric-power production and marketing. But as a result of pressure from new constituencies, such as trout fishers and river rafters, and the mandates of the National Environmental Policy Act and the Endangered Species Act (ESA), BOR initiated the Grand Canyon Environmental Studies (GCES).

After initiating the GCES, the BOR decided that it needed outside help to gain credibility for the studies. The Research Council was asked to form a committee to provide scientific review of the GCES, which it did from 1986 to 1995. The committee interacted with the GCES, provided regular reviews and comments, and summarized its work in its 1996 report (NRC 1996:16–8). Early and major criticisms of the GCES effort was that it lacked a coherent ecosystem perspective and that its process for getting external advice was slanted toward the views of BOR and WAPA, even though BOR received 33,000 comments on its draft environmental impact statement (EIS) for operation of the dam. The GCES was also limited initially to considering effects only in the region immediately below the dam, although the region within which effects occurred was much broader. In part, that limited conception of the affected region was a result of law and politics, but the Research Council reviews helped to expand the region of concern to areas that could reasonably be expected to be affected by changes in the dam's operations (NRC 1996:28–33).

One result of the GCES was consideration by BOR of alternative ways of restoring some features of the downstream ecosystem through controlled releases of impounded water at the Glen Canyon Dam. A finding that the river's tributaries below the dam supplied enough sediments to rebuild sandbars but that the muted flows of water as a result of the dam did not suffice to lift the sediments

was a key piece of information. In effect, BOR used an adaptive management strategy that recognized the need to weigh power-production costs and benefits against environmental costs and benefits. It also agreed to an experimental release, which was delayed to March 1996, to test the results of the GCES (NRC 1996:219–20).

The Valuation Problem

As the GCES process developed, the matter of weighing conflicting values grew in importance. On the one side were the revenues from hydropower generation and, of course, the necessary commitment to supplying water to users. Although electric power is sold to users who pay in ordinary dollars, the valuation issue is not simple. The economics of the electric-power industry are changing, and prices in a regulated industry are not always a good measure of real economic values. On the other side were a variety of values that are even more difficult to measure: recreation, including trout fishing and rafting; maintaining biological communities, including endangered species; maintaining riparian communities, including geomorphic formations; providing nonuse values, such as scenic and aesthetic resources; and preserving sites of cultural or archeological importance, including those related to Indian cultures. The GCES addressed the costs of reducing power generation in considerable detail. Nevertheless, the Research Council report concluded that it was difficult to resolve all the problems in the power-economics analyses.

Distributional issues across an array of affected interests, such as those involving investor-owned and municipal utilities and cooperatives, also affected the perspective used in the analyses. The Research Council committee concluded that a national perspective, as opposed to a regional or industry perspective, is appropriate because the dam is federally owned and the resources affected are of national, and perhaps even international, importance (NRC 1996:181). Economics research methods were also used in an effort to estimate the effects, in monetary terms, of water-release alternatives on river recreation and nonuse values. National and regional economic effects of the alternatives on recreation were estimated. The largest changes in economic effects were those involving commercial white water boating. Estimates of impacts on nonuse values derived with a CV method were based on household surveys in the region and nationally. These estimates were compared with estimated values of power production and recreation for the alternative levels of water flow. The results indicated that the estimated national nonuse values overwhelmed the estimated power, recreation, and regional nonuse values in evaluating river-flow alternatives (NRC 1996:118–36).

Results—The Grand Flush

For 1 week at the end of March 1996, BOR conducted its experiment by releasing 45,000 ft 3 of water per second—about 4 times the normal flow through Glen Canyon Dam. The results were visible within a day as new sediments added 0.5–3.0 m to the height of sandbars and beaches downstream. The results were almost immediately called a success with respect to improving recreational conditions throughout the Grand Canyon area and improving habitat for the humpback chub (NY Times 1996; Shapard 1996). At the same time, officials noted that this did not necessarily mean that such releases would become the norm, either at Glen Canyon or at other dams around the country. The estimated loss of power revenues of $2.7 million was not the most important factor in the assessment. Rather, more time is needed for a complete assessment of the environmental effects. Future large releases at Glen Canyon Dam must still fit with the need to provide water and power for traditional users (NY Times 1996; Shapard 1996)

An issue to be faced in the future is the possible effects of large water releases from Glen Canyon Dam on an endangered species, the Kanab amber snail, whose habitat is some 75 km below the dam. According to a Fish and Wildlife Service official, a third population of these snails will have to be established in addition to the two existing populations before another flood of this magnitude is permitted. Thus, whatever the other costs and benefits of occasional planned flooding of the river, they can be overridden by the requirements of the Endangered Species Act.

The experimental release of a surge of water from the Glen Canyon Dam had the expected beneficial environmental effects throughout the Grand Canyon. The costs in lost power generation appear to have been within reasonable bounds. The 10 years of studies that preceded the release, which cost about $50 million, received continuing scientific guidance from a Research Council committee, laid a solid foundation for the decision to go ahead with the experiment. Of particular importance was the finding during the course of the study that even though the Glen Canyon Dam blocked the downstream flow of sediments from above the dam, there was sufficient input of sediments from tributaries below the dam to justify the belief that sandbars and beaches could be restored. That was an important scientific consideration in the decision to go ahead with the experiment.

Although meeting biodiversity and other environmental objectives was a major goal of the experiment, it was also constrained by the legal mandates for the dam, which emphasized providing storage to regulate water flows to meet needs of users and to provide electric power at a reasonable price. The decision to go ahead with the experiment appear to have been made by responsible officials with detailed advice from scientists. Although it is clear that careful consid-

eration was given to the wide range of interests in the results of the test, including the varied interests of the Indian tribes along the Grand Canyon, the process used in getting inputs from these interests was not clear in the reports on which this case study is based.

Applicability to biodiversity

What is different about biodiversity (from the perspective of valuation).

Biodiversity presents two kinds of challenges for valuation: complexity and preponderance of nonmarket benefits. Biodiversity is an especially complex good, service, or amenity; in some ways, it is all these things. It might be most appropriate to think of biodiversity as a state of the world. Biodiversity involves complex suites of environmental services and amenities, which raise difficulties of several kinds:

• Information needs concerning the productivity of the environment with and without the proposed action in terms of services and amenities, stretch the capacity of the natural sciences.
• Many of the environmental services and amenities are unfamiliar to ordinary citizens, whose WTP and WTA are the fundamental information for valuation.
• Theoretical requirements for valuation of complex policies are systematically violated by schemes that value components of biodiversity separately, each by whatever method is feasible and appropriate, and then calculate total economic value by adding the component values. Valid approaches are limited to holistic total value (for example, one-shot WTP for the proposed change in the state of the world, with respect to the particular habitat) and sequential piecewise valuation in which each successive component is valued, assuming that budgets have been adjusted for WTP for all components earlier in the sequence (Hoehn and Randall 1987). Perhaps unsurprisingly, sequential piecewise valuation has seldom been implemented, and published valuation efforts using the piecewise strategy typically are susceptible to criticism in this regard.

Although we should be mindful of the inherent complexity of biodiversity, not all valuation tasks call on us to address its full extent (see the Pacific Northwest case study, below). Often, we are asked to value not total biodiversity, but the benefits and costs of proposed actions that would make relatively modest changes in it, for example, actions addressing just a few species in a particular place or modifying a particular section of habitat. The challenge of valuing modest changes in biodiversity is not trivial, but the task is more manageable than valuing total biodiversity. We note that, according to Norton (1988), a long series of marginal decisions can collectively make nonmarginal changes; and WTP to prevent those nonmarginal changes is greater than the sum of the mar-

ginal WTPs to make the marginal changes. So we could wake up one day to find that, through a series of small and individually rational decisions, we have "sold off" too much biodiversity too cheaply.

The second kind of challenge that biodiversity presents for valuation is the preponderance of nonmarket benefits—the combination (as befits a state of the world) of passive-use values, nonmarket-use values, and values arising from uncertain future uses of various kinds; the latter presents the greatest challenges for valuation. Accordingly, the role of market-oriented valuation methods can be relatively limited. Furthermore, if a holistic approach to total economic value is taken, that would tilt the choice of valuation method toward CV.

Some of the problems of weighing nonmarket benefits and costs are shown in the case study of the Pacific Northwest forests. A large part of the old-growth federal forests in the Pacific Northwest was reserved to protect habitat for the northern spotted owl and other species only after President Clinton intervened. The supporting analysis showed some—but only some—of the incremental costs of providing additional levels of habitat protection, but it was not possible to assign comparable estimates to the value of habitat protection. Despite the limitations of the analysis, it was helpful in weighing the alternatives considered in this presidential decision that involved aspects of biodiversity.

Case Study: Pacific Northwest Forests

This case involves a conflict between protection of species that depend on old-growth forests and long-standing use of the Pacific Northwest's forests as a major source of timber and wood products for national and international markets. Although the conflict started over protection of a single endangered species, the northern spotted owl, it grew into an issue involving protection of a suite of species, including various strains of salmon. At stake in the resolution of this conflict was not only the protection of species, but also sustaining the ecosystems that produce timber products and recreation services and support the people and communities that depend on the region's forests.

The northern spotted owl issue in the Pacific Northwest started in the late 1970s when biologists began to express concern over the loss of old-growth forests, which were believed to be the primary nesting, roosting, and foraging habitat for the owl. The reduction in area of the original old-growth forests was greatest in private forests, which had long been a major source of forest products. But the concern over loss of old-growth habitat focused on federal forests, which by the 1970s contained most of the remaining old growth and had also become an important source of forest products.

The spotted owl was listed as an endangered species under the ESA, and several reports confirmed that old-growth forests provided important habitat. Old-growth forests were also found to provide important habitat for the marbled murrelet and to help protect spawning habitat for salmon. In response to ESA

suits brought by environmental groups, a federal court halted further harvesting of old-growth timber on federal lands in the region until a satisfactory EIS had been prepared and actions proposed that would meet the requirements of the ESA. Logging on federal forests in the region was brought to a virtual standstill.

In an effort to find a solution, President Clinton convened a 1-day conference in Portland, Oregon, that soon led to yet another report, the report of the Forest Ecosystem Management Assessment Team (FEMAT). The FEMAT report and accompanying EIS present a detailed analysis of this federal public-lands issue involving biodiversity (FEMAT 1993).

Ten alternative policy options for management of federal lands within the northern spotted owl region were defined and compared. The president chose an option that created a system of reserved late-successional forest areas, riparian reserves, and so-called adaptive-management areas. The reserves more than doubled the area that was already congressionally reserved but left some federal forest available for a mix of resource uses. The president's choice was driven by the need to provide for the future viability of listed endangered species, especially the northern spotted owl and the marbled murrelet, to meet requirements set by the court.

The options differed in estimates of the future likelihood of ''habitat outcomes'' related to the expected viability of various species that are believed to depend on the continued existence of old-growth forests in the northern spotted owl region and estimates of expected effects on timber harvests and associated employment and economic measures. Expected habitat outcomes for each option were estimated separately for 48 species or groups of species of fungi, 16 groups of species of lichens, 12 species and 13 groups of species of bryophytes, 131 species of vascular plants, 102 species of mollusks, 15 species groups of arthropods, 18 species of amphibians, 15 species of mammals, 11 species of bats, 7 groups of species or races of fishes, and 38 species of birds, including the northern spotted owl and the marbled murrelet. The likelihood of projected future habitat outcomes for each species or group of species for each of the policy options was estimated for four categories: well distributed; locally restricted; restricted to refugia; and extirpation. The expected outcomes were estimated by a small group of specialists for each major group of species (fungi, lichens, bryophytes, and so on who were, in effect, required to agree on the estimates of the likelihoods for the four categories of habitat outcomes.

Other groups of specialists provided estimates of other outcomes for each policy option, including some probable annual volumes of federal timber harvests, direct timber-industry employment, and annual federal receipts from the sale of federal timber, all within the defined range of the northern spotted owl. These, of course, are only partial measures of economics-related effects. Several other kinds of economic-related impacts were discussed in the report, including the outlook for production of commodities other than timber, such as minerals; forage for range livestock; production of "special" forest products, such as floral

greens and wild edible mushrooms; outdoor recreation; scenic, water, and air quality; and other "public goods". For all the nontimber outputs, however, the specialists found that it was not possible to provide quantitative estimates of levels of production or output for each policy option.

Results for three of the nine policy options illustrate incremental changes in estimated habitat outcomes and in estimated economics-related measures between one policy option and another. Three policy options, in increasing order of degree of protection afforded to habitat for the various groups of species identified above, are

• No action—based on the federal lands management direction in place in 1992.
• President's choice—adopted by President Clinton after the spring 1993 meeting in Portland.
• Environmental option—would have put the largest area of federal land in protected reserves.

A comparison of the three options, their effects on three economics-related measures, and their effects on the availability of "well-distributed" habitat for the northern spotted owl, the marbled murrelet and coho salmon (three of the 426 species or groups of species for which estimates of the likelihood of habitat outcomes were made) is presented in table 5-1 .

The table shows the changes in the likelihood of well-distributed habitat for each of three species and the associated decreases in three economics-related measures. Obviously, the president's choice as the "best" overall alternative management regime for federal forests in the spotted owl region does not maximize favorable habitat outcomes. The choice was based on weighing the costs of mainly economics-related impacts against the benefits of added species protection.

For some species, such as the marbled murrelet and coho salmon, the estimated improvements in expected habitat conditions from adopting the president's choice are fairly dramatic relative to staying with the 1992 management regime (for example, from 26% to 80% for the marbled murrelet). But for the northern spotted owl, the improvement is less dramatic. The estimated additional improvements in habitat conditions from adopting the environmental option are not as large, and the estimated costs in terms of decreased federal timber harvests, and federal timber-sales receipts are quite large. The expected decreases in timber-related employment are not proportional to the decreases in federal timber harvests, because harvests from federal lands make up only part of the total regional timber harvest.

The decision in this case was aimed not at protecting biodiversity itself, but at resolving an issue related to protecting endangered species. But the basis of the decision was an evaluation of habitat outcomes of a long list of species and a range of alternatives. The approach for estimating habitat outcomes was to force

TABLE 5-1 Comparison of the Three FEMAT Options

a small group of specialists in each category of species to come to an agreement in a short time with only the currently available information. Perhaps years of additional research would lead to different estimates, but resource managers usually cannot wait that long. The approach provided reasonably adequate information for recognizing real differences in habitat outcomes among the management options that were considered.

The economics-related measures of differences in effects among the options were also reasonable, although limited. The decision was political. Effects on employment and receipts from timber sales are relevant when federal forests are involved. But they are only partial measures of the values involved in maintaining some degree of biodiversity.

Effects on communities in the region and the value of interagency and citizen collaboration in making these kinds of decisions were also recognized in the FEMAT report. And there was a short discussion of the economic effects of increased prices (due to decreased federal timber harvests) on consumers of wood products. But estimates of the various effects with the different options were not included in the report, largely because of conceptual and measurement problems. There was no single "currency" with which the value of biodiversity could be measured for this resource management decision even if there had been clear agreement on how to measure differences in biodiversity.

The size of the region was based largely on the range of the northern spotted owl, which was the subject of the original court case. That meant it was a very large area for focusing on some kinds of effects, such as human-community impacts, which vary considerably from one locale to another whose measurement can easily be lost when impacts on individual communities are melded into state or multistate regional estimates. At the same time, once the issue was defined to include the impacts of forest management on salmon, it was difficult to leave out the effects on salmon of management of ocean fisheries, which extend the scale even more. In contrast, evaluation of habitat outcomes for some of the species that have restricted ranges, such as some of the mollusks, might have lost something by being part of a bigger analysis.

This case shows of the incremental effects of a resource-management decision involving biodiversity. The FEMAT analysis provided input for the decision by showing what would be gained and lost for each additional increment of protection of old-growth forest habitat. The basic structure of the analysis was appropriate even if the analysis was limited by gaps in available information.

Roles for the Various Valuation Methods

Direct market evidence can be useful for valuing natural-resource commodities harvested from biodiverse environments, genetic resources useful (for example) in plant breeding for agriculture and forestry, pharmaceutical resources, and so on.

The travel-cost method is useful for evaluating recreational-use benefits. The growing demand for adventure travel and ecotourism suggests that such benefits will play an increasing role, especially in habitats where charismatic megafauna are present and tourism can be managed compatibly with species and habitat preservation.

Hedonic price analysis can have a role in benefit estimation, for example, in cases where a market develops in land near habitat reserves so that land value reflects demands for amenities generated by biodiversity. Nevertheless, it is reasonable to expect the role of such analysis to be modest and occasional.

The methods relying on direct and indirect evidence from markets have important limitations for valuing biodiversity:

• They have some conceptual and methodological problems.
• They are limited to a subset of use values; passive-use value and holistic total economic value are beyond their reach.
• It is difficult to implement conceptually valid piecewise valuation procedures that would give these methods a role in eclectic valuation schemes that use different methods to value different components of the suite of biodiversity services and amenities.

Contingent valuation , although controversial, is the obvious method for valuing biodiversity because it is, at least in principle, capable of valuing nonmarket-use values, passive-use values, and total economic value. Nevertheless, biodiversity presents serious challenges for CV in that respondents often are asked (of necessity) to value relatively unfamiliar services and amenities.

Of the more than 2,000 publications to date involving CV, relatively few have addressed biodiversity, habitats, or endangered species. Passive-use values, because they are nonrival (that is, passive users do not compete with each other for access), can be very large in the case of environmental services that appeal to a large number of people. In a CV of viewing of elephants in Kenya, Brown and Henry (1993) elicited WTP for maintenance of elephant populations. Their results show that visitors gain about $25–30 million per year in consumer surplus (value over what they actually pay) for viewing elephants, a proportion of which is likely to represent passive use value. On wider habitat protection, Moran (1994) shows that the consumers' surplus attached to nonconsumptive use of Kenya's protected areas by foreign visitors (as a subset of all users) is about $450 million. It is safe to conjecture that those values would be overshadowed by the passive-use values of nonvisitors if these had been measured.

Given the broad applicability of CV to valuing biodiversity, it is important to address the criticisms that have been raised about CV:

• Validation of CV is inherently difficult. In the absence of convincing validation, and given the very large value estimates that can be expected for

passive-use value and total economic value of prominent biodiversity resources, CV and the value estimates that it generates will remain controversial.

• CV surveys of biodiversity or species-preservation issues often generate a relatively high proportion of protest or refusal responses, and some respondents indicate an unwillingness to address these issues in terms of trades for money. Good CV design—for example, structuring the CV as a referendum about spending more or less public money for preservation projects—can minimize the occurrence of protest or refusal responses. Nevertheless, some refuse to respond to even the best CV questions, and some of these nonrespondents are thoughtful people who draw on nonutilitarian moral philosophies when trying to resolve biodiversity issues. Chapter 4 makes clear that these are legitimate reactions, and they illustrate the limits of utilitarian CV in dealing with nonutilitarian concerns.
• CV comes in a variety of forms, each with its own communication and incentive properties, so blanket claims about the validity of CV are meaningless. But one constant is that the validity of any CV effort depends on respondents' understanding of what is being valued. In the case of biodiversity, citizen knowledge of the details of any particular case is likely to be quite low, so researchers will need to provide a good deal of case-specific information. Therefore, issues of communication and comprehension are likely to be prominent in criticism of many CV efforts directed at biodiversity. It is important to recognize that this problem applies also to any other approach or process that takes citizen opinion seriously.

Contingent-choice experiments are still in their infancy, especially in contexts where passive-use values can well dominate. Nevertheless, one might expect increasing application of these methods.

Various estimates have been made of the value of aspects of biodiversity. They include in this report the estimates for ecosystem services in chapter 3 , the estimates in the Pacific Northwest forests and Grand Canyon flush case studies in this chapter. In addition, the article by Costanza and others (1977) discussed in a later section, A Tempest Over Valuing the World's Ecosystem Services , has global estimates of average per hectare and total values of biodiversity for 17 ecosystem services and 17 biomes.

In most of the cited examples, as well as in most of the numerous other published examples, the value estimates are for some particular element of biodiversity or for services that are related to some element of biodiversity. The estimates in the paper by Costanza and others (1997) are unusual in that the sum of the values for the 17 ecosystem services represents estimates of one-time annual values or the present value of the stream of expected future values.

Benefit-Cost Analysis in the Federal Government

Benefit-cost analysis (BCA) is a generic term that can refer to nearly any comparison of benefits and costs as long as they are measured or estimated in comparable units. In the federal government, the term sometimes refers to protocols for comparing the "desirable and undesirable impacts of proposed policies" (see box 5-2 ) (Arrow and others 1996). An early use of a formal process of BCA was in the evaluation of federal water-resources development projects after enactment of the Flood Control Act of 1936. The act required that proposed projects be undertaken only "if the benefits to whomsoever they accrue exceed the costs." That, of course, is consistent with the progressive model of "scientific govern-

ment", a model that achieved great influence in the first half of this century (Nelson 1987).

Guidance for analyses of federal water-resources projects grew in a series of documents (the "Green Book" of 1947, Senate Resolution 148 of 1957, Senate Document 97 of 1962, and so on) that ultimately provide for a "four accounts" model for water-resources planning: the national economic-development account (basically, a BCA); the regional economic-development account, which focuses on income and employment effects; the environmental-quality account, which overlaps with the national economic-development account to the extent that it proves feasible to determine the benefits and costs of at least some of the antici-

pated environmental-quality changes; and the quality-of-life account. Both benefits and costs were to be measured in terms of estimated market values. The approach appears straightforward on the surface, but numerous issues arise in trying to apply the general guidelines to proposed projects, including defining the appropriate scope of the analysis, deciding how to account for regional and local effects, and especially deciding how to treat real benefits and costs for which market values do not exist.

Concerns in the last 2 decades with the presumed high costs of federal regulations has broadened the use of BCA by the federal government. Executive order 12291 of 1981 required analysis of benefits and costs of all new federal regulations with major economic and other effects. It stated that regulatory actions should maximize "net benefits to society" and should not be undertaken "unless the potential benefits to society . . . outweigh the potential costs to society." The order, taking note of the difficulty of estimating both benefits and costs in monetary terms, required that proposed rules include descriptions of benefits and costs that cannot be quantified and that the determination of net benefits include "an evaluation of the effects that cannot be quantified in monetary terms.'' Executive order 12866 of 1993 replaced executive order 12291. It, too, required that assessments of benefits and costs include qualitative measures of those which are difficult to quantify. A recent article by several prominent economists noted that estimates of benefits and costs of regulations should be accompanied by a description of the uncertainties surrounding the estimates and that the analyses should also identify distributional (equity) consequences (Arrow and others 1996). But the executive orders provide few clues about how these qualitative measures are to be made and used in analyses. BCA for federal regulations continues to be treated on a largely ad hoc and project-by-project basis, especially where the market provides little guidance on values.

With the fading of the progressive dream, a pluralistic, participatory process has emerged. Instead of trusting in the experts to get things right, citizens seek access to the decision-making process. BCA has a somewhat different role in such a process: it will have influence to the extent that citizens are convinced, first, that benefits and costs are relevant considerations and, second, that the particular BCA is reasonably accurate and reliable. The second of those concerns—essentially, the quality of benefit-cost information—is a serious concern in the context of biodiversity, but it has already been discussed here at some length. Here, we address the first concern: Can we give good reasons why benefits and costs are relevant considerations in policy decisions?

• A BCA is an account—not a perfect account but a fairly good account—of the prospective contribution of some proposed action to the satisfaction of human preferences. Because preference satisfaction cannot logically count for everything but also cannot logically count for nothing, benefits and costs will be

relevant considerations in policy decisions but will not be the only relevant considerations (Hubin 1994; Randall 1999).

One appealing answer to the question of when BCA counts and how much it counts is that an ethically pluralistic society of thoughtful moral agents would have good reasons to agree to choose on the basis of benefit-cost considerations when nothing more important is at stake and to impose the most important of these "more important" considerations as constraints on what can be chosen (Randall 1999).

• In the case of biodiversity, the appropriate constraints can well include safe minimum standards of conservation for critical species and habitats (see box 5-3 ) (Farmer and Randall 1998).

A Tempest over Valuing the World's Ecosystem Services

In May 1997, Robert Costanza and a long list of coauthors published a paper titled "The Value of the World's Ecosystem Services and Natural Capital" in Nature . They estimate the annual value of the world's ecosystem services to be about $36 trillion, compared with an estimate of about $18 trillion for the world's annual gross product. The reaction of neoclassical economists is best typified by V. Kerry Smith's response (Mispriced Planet, Summer 1997, Regulation ): "Their results should not be used in any form—whether as measures of the importance of changes in natural resources to human welfare; as yardsticks for future project appraisals; or as sources of a road map for future research." Although there are many technical issues for debate, the core of the argument arises because of the claim by Costanza and others (1997) that their estimate of the value of ecosystem services is based on the concept of WTP.

A thought experiment is useful. Imagine that evil aliens orbit Earth and threaten to destroy ecosystem resources one by one unless we pay blackmail in the form of an annual fee for each service. Costanza and his colleagues are quickly assembled to value each category of ecosystem services. The first resource threatened is forests, which generate $4.7 trillion per year, on the basis of the estimated WTP of the world's countries for the forests' total ecosystem services (Costanza and others 1997). On the basis of the group's recommendations, Earth agrees to pay $4.7 trillion each year. Next the aliens threaten the coastal shelves, worth $4.3 trillion. However, having already agreed to pay $4.7 trillion for forests, reducing available world gross product for human consumption from $18 to $13.3 trillion, Earth opposes the Costanza estimates because "we cannot afford $4.3 trillion more; we are much poorer now." In other words, the demand and value for coastal shelves is reduced because available gross product from which to pay is reduced. If we follow this line of argument, the world's total annual gross product ($18 trillion) is the most that could be paid as a bribe to save the world's ecosystem services without reducing the accumulated value of the the world's capital.

The value estimates of Costanza and others (1997) are based on separate studies of the values of individual components, each of which assumes that people's incomes remain at current levels. The problem has been termed the independent valuation and summation problem by Hoehn and Randall (1989), who argue that it is inappropriate to simply add the values obtained from independent studies, because aggregate values will be overstated. It is clear from the way that Costanza and others construct their estimates that their work does, in fact, suffer from the independent valuation and summation problem. However, the story is not over.

Costanza and others (1997) respond to Smith with a substantive counter-argument of their own. Because ecosystem services are, for the most part,

unpriced, the sum of the world's gross product underestimates world income. Furthermore, the actual value of the world's ecosystem services would increase through proper management if the resources were properly priced. A simple example will illustrate their argument: Because of overfishing, the North Atlantic fishery is now capable of contributing little to the world's gross product. With proper management (which might include putting a price or tax on each fish taken from the sea to discourage overfishing), the fishery would be restored, the sum of the world's gross product would go up, and our ability to pay a bribe to protect coastal shelves from alien destruction would increase.

The utility of the paper by Costanza and others (1997) is not in its estimates of the value of the world's ecosystem services, but rather that it initiated a visible discussion of the difficulties of estimating such values, whether on a global or on a more localized basis. As was pointed out in chapter 2 , biological systems are complex. The debates over the Costanza paper point to the complexity and interactions of economic systems. These debates have contributed to a better public understanding of the difficulties in estimating economic values, especially in the absence of market-price information. As long as the value of most ecosystem services is not subjected to a market test, such debates will continue, and in the end they will advance understanding not only of the issues, but also of the values that are involved.

Economists have developed an array of tools for estimating values when the lack of ordinary markets precludes use of their favored measure, market-determined prices. These are powerful tools for informing decisions involving biodiversity. But they have limitations. Estimates of value based on them should be treated with careful attention to the assumptions that have been made in obtaining them. Support for their veracity can be indicated by the degree to which results obtained from various estimates converge. Particular care should be taken as the scale of the decisions for which estimates of value are made diverges from the normal scale of market processes. The economist's usual view of market decisions as being made at the margin—that is, for small changes in quantities and prices—is a key assumption for most estimates of value.

Adamowicz WL, Louviere J, Williams M. 1994. Combining revealed and stated preference methods for valuing environmental amenities. J Envir Econ Mgmt 26:271–92.

Arrow KJ, Solow R, Portney PR, Leamer EE, Radner R, Schuman J. 1993. Report of the NOAA panel on contingent valuation. Washington DC: GPO.

Arrow KJ, Cropper ML, Eads GC, Hahn RW, Lave LB, Noll RG, Portney PR, Russell M, Schmalensee R, Smith VK, Stavins RN. 1996. Is there a role for benefit-cost analysis in environmental, health, and safety regulation? Science 272(5259): 221–2.

Bishop RC. 1978. Economics of endangered species. Amer J Agric Econ 60:10–18.

Bockstael N. 1995. Travel cost models. In: Bromley DW (ed). The handbook of environmental economics. Cambridge MA: Basil Blackwell. p 655–71.

Bockstael N, Hanemann ME, Kling CL. 1987. Estimating the value of water quality improvements in a recreational demand framework. Water Resour Res 23:951–60.

Brown GM, Henry W. 1993. The economic value of elephants. In: Barbier EB (ed). Economics and ecology: new frontiers and sustainable development. London UK: Chapman & Hall.

Carson R, Flores N, Martin K, Wright J. 1996. Contingent valuation and revealed preference methodologies: comparing the estimates for quasi-public goods. Land Econ 72:80–99.

Ciriacy-Wantrup, S von. 1968. Resource conservation: economics and policies, 3rd ed. Berkeley CA: Univ Calif Div Agric Sci.

Costanza R, d'Arge R, de Groot R, Farber S, Grasso M, Hannon B, Limburg K, Naeem S, O'Neill RV, Pareuelo J, Raskin RG, Sutton P, van den Belt M. 1996. The value of the world's ecosystem services and natural capital. Nature 387:253–60.

Daly HE, Cobb JBJ. 1989. For the common good: redirecting the economy toward community, the environment, and a sustainable future. New York NY: Beacon City Pr.

Farmer M, Randall A. 1997. Policies for sustainability: lessons from an overlapping generations model. Land Econ 3:608–22.

Farmer M, Randall A. 1998. The rationality of a safe minimum standard of conservation. Land Econ 74:287–302.

FEMAT [Forest Ecosystem Management Assessment Team]. 1993. Forest ecosystem management: an ecological, economic, and social assessment. Washington DC: USDA, Forest Service, and USDOI, Fish and Wildlife Service, and others.

Hanemann WM. 1984. Welfare evaluations in contingent valuation experiments with discrete responses. Amer J Agric Econ 66:332–41.

Hausman J. 1993. Contingent valuation: a critical assessment. Amsterdam Netherlands: Elsevier Science.

Heywood CH (ed). 1995. Global biodiversity assessment. Cambridge UK: Cambridge Univ Pr.

Hoehn J, Randall A. 1989. Too many projects pass the benefit cost test. Amer Econ Rev 79:544–51

Hoehn J, Randall A. 1987. A satisfactory benefit cost indicator from contingent valuation. J Envir Econ Mgmt 14:226–47.

Howarth RB, Norgaard RB. 1990. Intergenerational resource rights, efficiency, and social optimality. Land Econ 66:1–11.

Hubin DC. 1994. The moral justification of benefit/cost analysis. Econ Phil 10:169–94.

Kant I. 1991. Philosophical writings. New York NY: Continuum. (Behler E [ed]).

Moran D. 1994. Contingent valuation and biodiversity conservation in Kenyan protected areas. Biod Cons 3(8): 663–84.

NRC [National Research Council]. 1996. River resource management in the Grand Canyon. Washington DC: National Acad Pr. 226 p.

NRC [National Research Council]. 1997. Valuing ground water: economic concepts and approaches. Washington DC: National Acad Pr.

Nelson RH. 1987. The economics profession and the making of public policy. J Econ Lit 25:49–91.

Norton B. 1988. Commodity, amenity, and morality: the limits of quantification in valuing biodiversity. In: Wilson EO, Peters FM (eds). Biodiversity. Washington DC: National Acad Pr. p 200–205.

Page T. 1977. Conservation and economic efficiency. Baltimore MD: Johns Hopkins Univ Pr.

Randall A. 1999. Taking benefits and costs seriously. In: Tietenberg T, Folmer H (eds). The international yearbook of environmental and resource economics. Cheltenham UK, Northampton MA: Edward Elgar.

Randall A. 1994. A difficulty with the travel cost method. Land Econ 70:88–96.

Randall A. 1987. Total economic value as a basis for policy. Trans Amer Fisheries Soc 116:325–35.

Shapard, R. 1996. A grand experiment brings spring floods to the canyon. Amer City Country 111:26

Smith VK, Osborne LL. 1996. Do contingent valuation estimates pass a Scope test? A meta analysis. J Envir Econ Mgmt 31:287–301.

Resource-management decisions, especially in the area of protecting and maintaining biodiversity, are usually incremental, limited in time by the ability to forecast conditions and human needs, and the result of tradeoffs between conservation and other management goals. The individual decisions may not have a major effect but can have a cumulative major effect.

Perspectives on Biodiversity reviews current understanding of the value of biodiversity and the methods that are useful in assessing that value in particular circumstances. It recommends and details a list of components—including diversity of species, genetic variability within and among species, distribution of species across the ecosystem, the aesthetic satisfaction derived from diversity, and the duty to preserve and protect biodiversity.

The book also recommends that more information about the role of biodiversity in sustaining natural resources be gathered and summarized in ways useful to managers. Acknowledging that decisions about biodiversity are necessarily qualitative and change over time because of the nonmarket nature of so many of the values, the committee recommends periodic reviews of management decisions.

READ FREE ONLINE

Welcome to OpenBook!

You're looking at OpenBook, NAP.edu's online reading room since 1999. Based on feedback from you, our users, we've made some improvements that make it easier than ever to read thousands of publications on our website.

Do you want to take a quick tour of the OpenBook's features?

Show this book's table of contents , where you can jump to any chapter by name.

...or use these buttons to go back to the previous chapter or skip to the next one.

Jump up to the previous page or down to the next one. Also, you can type in a page number and press Enter to go directly to that page in the book.

Switch between the Original Pages , where you can read the report as it appeared in print, and Text Pages for the web version, where you can highlight and search the text.

To search the entire text of this book, type in your search term here and press Enter .

Share a link to this book page on your preferred social network or via email.

View our suggested citation for this chapter.

Ready to take your reading offline? Click here to buy this book in print or download it as a free PDF, if available.

Get Email Updates

Do you enjoy reading reports from the Academies online for free ? Sign up for email notifications and we'll let you know about new publications in your areas of interest when they're released.

An official website of the United States government

The .gov means it’s official. Federal government websites often end in .gov or .mil. Before sharing sensitive information, make sure you’re on a federal government site.

The site is secure. The https:// ensures that you are connecting to the official website and that any information you provide is encrypted and transmitted securely.

• Publications
• Account settings

Preview improvements coming to the PMC website in October 2024. Learn More or Try it out now .

• Advanced Search
• Journal List
• Int J Environ Res Public Health

Analysis of Academic Literature on Environmental Valuation

Francisco guijarro.

1 Research Institute for Pure and Applied Mathematics, Universitat Politècnica de València, 46022 Valencia, Spain

Prodromos Tsinaslanidis

2 Department of Economics, University of Western Macedonia, 52100 Kastoria, Greece

Environmental valuation refers to a variety of techniques to assign monetary values to environmental impacts, especially non-market impacts. It has experienced a steady growth in the number of publications on the subject in the last 30 years. We performed a search for papers containing the term “environmental valuation” in the title, abstract, or keywords. The search was conducted with an online literature search engine of the Web of Science (WoS) electronic databases. A search of this database revealed that the term “environmental valuation” appeared for the first time in 1987. Since then a large number of studies have been published, including significant breakthroughs in theory and applications. In the present work 661 publications were selected for a review of the literature on environmental valuation over the period 1987–2019. This paper analyzes the evolution of the leading methodologies and authors, highlights the preference for the choice experiment method over the contingent valuation method, and shows that relatively few papers have had a strong impact on the researchers in this area.

1. Introduction

Environmental valuation has traditionally been considered in the context of non-market valuation. Its aim is to obtain a monetary measure of the benefit or cost to the welfare of individuals and social groups of environmental improvement interventions or the consequences of environmental degradation [ 1 , 2 ]. However, the ultimate goal is not to value a (non-market) environmental good in monetary terms, but to provide decision-makers with the necessary tools to take the appropriate political initiatives to efficiently allocate resources, impose taxes and design compensation schemes [ 3 , 4 ], even after assuming the difficulties of developing theoretically grounded practical policy tools and avoiding political manipulation [ 5 ].

Environmental valuation methods have been used to determine the benefits and costs related to the use of environmental goods, improving their conditions or remedying environmental damage and must consider the complexity of the area. For example, the economic benefits of national parks extend beyond tourism; natural amenities and recreation facilities often serve to attract and retain people, entrepreneurs, businesses, and retirees [ 6 ]. On the other hand, some researchers have provided evidence of how worsening environmental conditions can affect the value of other goods. For example, noise and air pollution from road traffic have been reported to negatively impact real estate prices [ 7 , 8 ], and [ 9 ] reported that 55% of those surveyed in Brisbane (Australia) considered that noise adversely affected the value of their property.

Economists have traditionally developed tools to measure environmental values by estimating individuals’ willingness to pay to benefit from environmental goods. The costs associated with environmental deterioration are measured by the loss suffered by the individuals who benefited from the damaged good, and deciding the appropriate compensation for losing the benefit (willingness to accept) [ 10 , 11 ].

The general approach of Total Economic Value (TEV) combines all the different values, which are grouped according to the service provided by the environmental good ( Figure 1 ). The use values are those derived from the actual use of the resource, while the non-use values are not related to its present use. The former includes the direct use value—the value derived from the direct use and exploitation of the environmental good, the ecological value—defined by the benefits that environmental goods provide to support forms of life and biodiversity and the option value—related to future use opportunities of the good. Non-use values are composed of the existence value—the value that individuals give to environmental goods for their mere existence—and the bequest value—the value estimated by individuals when considering the use of goods in the future by their heirs.

The concept of Total Economic Value of environment, taken from [ 12 ].

The aim of environmental valuation methods is to measure the values included in TEV. Although some authors have classified valuation methods from a more general perspective [ 13 ], the methods specifically related to environmental valuation can be classified as follows:

• - Contingent valuation method. Values are estimated in a hypothetical market based on surveys in which respondents are asked how much they are willing to pay for the use and conservation of an environmental good. The purpose of contingent valuation is to estimate individual willingness to pay for changes in the quantity or quality of environmental goods or services [ 3 ].
• - Choice experiment method. This method provides the respondents with alternative choices in which different environmental goods are defined by their attributes. According to [ 14 ], “the most significant advance in environmental valuation may be to move away from a focus on value and focus instead on choice behaviour and data that generate information on choices.”.
• - Travel cost method. Values are estimated by accounting for the cost incurred by people who travel to visit an environmental good. The method assumes that the willingness to pay must be at least as large as the travel cost incurred.
• - Hedonic price method. Values are computed from the prices of traded goods. This approach is frequently used when the price of traded goods is influenced by environmental factors [ 8 ].

The field of environmental valuation has recently expanded both from a theoretical and practical point of view [ 15 ]. This paper aims to outline the advances made by researchers according to their impact on the research area and highlights the key aspects covered by leaders in this field.

To determine the most important topics and assess the academic impact of environmental valuation, we performed a bibliometric analysis considering publications in the Web of Science from 1987 to 2019. We assessed their productivity through their historical evolution and the distribution of papers by journal. The units of analysis were ordered by the citation and co-citation structure and the results gave insights into the organization and future trends on research in environmental valuation.

We performed a search for papers containing the term “environmental valuation” in the title, abstract or keywords. The search was conducted on the online literature search engine of the Web of Science electronic databases. On 17 December 2019 we obtained 661 results from the search engine covering the period 1987–2019, including articles, book chapters, proceedings papers and reviews of 1442 authors. Table 1 shows the protocol followed to perform the data collection and some key figures.

Procedure for the data collection and key figures.

The dataset is analysed in the following section on R [ 16 ], a free software environment for statistical computing and graphics. The bibliometrix [ 17 ] package was used to compile most of the tables in this paper.

3.1. Environmental Evaluation Publication History

The number of publications per year is depicted in Figure 2 . The first known paper on environmental valuation, published in 1987, was followed by a steady increase in number of environmental valuation-related publications over time.

Distribution of Environmental valuation publications by year (1987–2019).

Although the research was published in a wide range of journals, the 4 most popular were: Ecological Economics, Environmental & Resource Economics, Environmental Values, and Journal of Environmental Management– with nearly 30% of the studies ( Table 2 ). Ecological Economics stands out as the most prolific source on this subject with 109 papers, which represents 16.5% of the total sample. Not surprisingly, the top Journals are particularly involved with environmental and ecological issues. The first 6 Journals are grouped into Environmental Sciences or Environmental Studies categories from the Journal Citation Reports of the Web of Science. When taking the impact factor into consideration, the top 6 Journals were ranked into the first quartile of their corresponding categories in 2018, while the rest of Journals are between the first and second quartile in most cases.

Most relevant journals that have published the greatest number of environmental valuation papers.

3.2. Leading Topics in Environmental Valuation Research

The most common keywords used by researchers include “environmental evaluation”, “willingness to pay”, and “ecosystem services” ( Table 3 ). The keyword “environmental valuation” was used in 38% of the publications analyzed. The following keywords give useful insights into the evolution of the research topic and the methods developed and applied to value environmental goods and damage. The second most often used keyword is “willingness to pay”, which is commonly found in publications related to stated preference methods. The two abovementioned approaches to this group of environmental valuation methods occupy positions 4 (choice experiment) and 5 (contingent valuation). The choice experiment method also appears in the 7th position as “choice experiments”. The total of both alternatives (78) comes just after the “environmental valuation” keyword.

The 10 most used keywords by number of publications related with environmental valuation.

As the search procedure is automatic, the system differentiates “Choice experiment” from “Choice experiments”. In order to consider all the possible synonyms, we conducted a new experiment by searching for individual terms in the keywords ( Table 3 ). For example, the word “choice” was used to collect all the papers with a keyword related to the choice experiment method. This provided similar expressions to those given in Table 3 : Choice modeling, Choice modelling, Choice model, Choice experiment method, etc. The analysis showed that keywords related to the choice experiment method appeared in 165 papers, while other methods had a lower frequency (contingent valuation method, 69; hedonic price method, 18; travel cost method, 11).

The relevance of choice experiments as a prominent keyword used by researchers has increased over time. We show the evolution of four keyword categories through 3 equally spaced subperiods: 1987–1997, 1998–2008 and 2009–2019 ( Figure 3 ). The first two subperiods were dominated by keywords associated with contingent valuation methods (with labels “contingent valuation” and “contingent valuation method”) and cost-benefit analysis. However, a sudden change was found in the trend during the subperiod 2009–2019. During this time the choice experiments (with labels “choice experiment”, “choice experiments” and “choice experiment model”) dominated the researchers’ interest, closely followed by the willingness to pay keyword.

Evolution of the main keywords used by researchers in environmental valuation.

The popularity of the choice experiment method –over the contingent valuation method—was predicted by Adamowicz [ 14 ]: “The most significant advance in environmental valuation may be to move away from a focus on value and focus instead on choice behaviour and data that generate information on choices.” We can suggest several reasons to support the observed trend. First, the design of both methodologies makes the choice experiment method to extract more information than the contingent valuation method does. Results from contingent valuation are elicited by asking respondents for their willingness to pay (or willingness to accept). In a bidding game, the respondent is asked if he is willing to pay a specific amount of money. If the answer is yes, a higher amount is asked and, if the answer is no, a lower amount is proposed. The questionnaire is repeated until an initial yes changes to a no or vice versa. However, the choice experiment method uses attributes to define alternatives and information of the willingness to pay is obtained by observing the choices made by respondents [ 18 ]. As stated by Hoyos [ 15 ], the choice experiment method allows estimating the mean willingness to pay and also the marginal willingness to pay for the different attributes. Handling with more alternatives and attributes makes the application of the choice experiment more complex. However, its implementation has been facilitated by the development of statistical software. Furthermore, web-based surveys are becoming popular and easy to implement and the number of connected people to the internet keeps increasing, which limits biased sampling, then allowing presenting the choice set in a friendly manner [ 19 ]. An additional benefit from using the choice experiment method is related with the sensitivity to scope. This is one of the main concerns about the contingent valuation method, where the use of labels in the choice experiment may mitigate the lack of sensitivity to the scope [ 19 ].

3.3. The Most Influential Authors in Environmental Valuation

The most prominent authors in an area of research can be identified by citation analysis. Of the top 10 most influential publications on environmental valuation according to the number of citations, Boxall and Adamowicz [ 20 ] leads with 527 citations ( Table 4 ). The authors use a latent class model to evaluate choice behaviour as a function of observable attributes of the choices and latent heterogeneity in the respondents’ characteristics. Although it has the highest number of citations, the paper by Lancsar and Louviere [ 21 ] received more cites on a yearly basis. The choice experiment model dominates the top ranked papers of Table 4 in which the authors introduce different environmental valuation examples to illustrate their proposals. Some of the top ranked papers are devoted either to the demonstration of case studies or to a review of the literature.

The 10 most frequently cited papers on environmental valuation.

We have analyzed the relevance of different authors in the topic according to the number of publications and the number of citations per year. Figure 4 gives one line to each author, where the extremes represent the year of the first (left circle) and last publication (right circle). Hanley was cited for the longest period, which was 25 years (1995–2019). The diameter of the circles varies in proportion to the number of papers published each year and the colour denotes the number of cites received. For example, the paper by Hanley et al. [ 22 ] has the highest number of citations per year (21.9) in the table. Although this is not the most cited paper according to the bibliographic analysis, it appears in the figure because Hanley is the most prolific researcher.

Relevance of authors according to their production and the number of citations.

The figure distinguishes two groups of authors. The first incorporates those who have been publishing on the topic for roughly 20 years: Hanley, Adamowicz, Boxall, Spash and Brouwer. The other group contains those who published between 2007 and 2019: Meyerhoff, Schaafsma, Hoyos, Mariel and Thorsen.

It should be noted that a few papers are responsible for a high percentage of the citations ( Figure 5 ). gives the number of citations in descending order. Only 7 papers received more than 300 citations for the whole period analyzed, while 55.7% received 10 or fewer. This shows that only a few papers influenced this research topic during this period.

Distribution of citations per paper.

Lastly, there is another interesting point related to the authors’ affiliation country; Figure 6 separates the papers whose authors’ affiliations are all located at the same country (Single Country Publications, SCP) and those with authors’ affiliations from different countries (Multiple Country Publications, MCP). The UK and the USA dominate the research on environmental valuation according to the number of papers published during the analyzed period. There are only 5 European countries in the top 10, while China is the only Asian representative. China is also in the last position in the top 10. Regarding the collaboration between authors from different countries, researchers from the UK and Spain are the most likely to collaborate in multinational publications, while Brazilian and Chinese affiliations produced the fewest publications with contributions from foreign authors.

Most productive countries in environmental valuation.

3.4. Co-Citation Analysis

This subsection begins with some comments about what productivity is in the field of research publication. Of course this is a wide field of debate, but some preliminaries must be established before proceeding with the co-citation analysis. According to [ 29 ], there are several measures to account for productivity. The most basic bibliometric measure is the number of papers published, which provides the raw data for all citation analysis. Another measure is the number of citations, which determines the recognition and influence of a paper. Then we can distinguish between citations received from papers published in Journals indexed in WoS, or citations received for other Journals not considered in WoS. As stated by [ 29 ], a measure of association between highly cited papers is used to form clusters: “That measure is the number of times pairs of papers have been co-cited, that is, the number of later papers that have cited both of them”. Hence, co-citation implies that two papers are cited in a third paper and assumes that both papers are related. We have performed a co-citation analysis by differentiating 3 main clusters in different colours ( Figure 7 ). The references of cluster 1 are represented by the book by Mitchell and Carson [ 30 ], in which the authors describe the contingent valuation method and claim that “the contingent valuation (CV) method offers the most promising approach for determining public willingness to pay for many public goods”. However, the positivist perspective in Mitchell and Carson [ 30 ] is contested by other prominent works in the same group. The report in Arrow et al. [ 31 ] indicate several drawbacks to the contingent valuation method and gives some guidelines to be used if the proposal is to produce useful information for natural resource damage assessment. The research in Kahneman and Knetsch [ 32 ] reports the most serious shortcoming of the CV method. According to these authors: “the assessed value of a public good is demonstrably arbitrary, because willingness to pay for the same good can vary over a wide range depending on whether the good is assessed on its own or embedded as part of a more inclusive package”. There is a more recent relevant book in this group, Bateman et al. [ 33 ], which gives a general approach to stated preferences techniques with application to different non-market goods and services.

Co-citation network analysis.

The cluster 2 (in red) elicited from the co-citation analysis is led by the paper by Boxall et al. [ 26 ], “A comparison of stated preference methods for environmental valuation”. This paper introduces an empirical comparison of the contingent valuation method and choice experiments. Most papers in this group follow the approach in Boxall et al. [ 26 ]. For example, Adamowicz et al. [ 34 ] examine the choice experiment as “an extension or variant of contingent valuation”. The paper in Adamowicz et al. [ 35 ] had earlier compared a stated preference model and a revealed preference model for recreational site choice. The earliest work in the group is the book by Ben-Akiva et al. [ 36 ], which analyzes the discrete choice method from a more general perspective.

And lastly, the cluster 3 covers different references related to choice modelling approaches but with a different approach to the publications in the second group. Again, a single book is the leader in number of cites: Louviere et al. [ 37 ]. Interestingly, this book is not the only reference which gives a survey of choice modelling. The paper by Hoyos [ 15 ] provides a review of the state of the art of environmental valuation with discrete choice experiments; Hanley et al. [ 22 ] examine the choice modelling approach to environmental valuation. The authors state that this methodology “can be considered as an alternative to more familiar valuation techniques based on stated preferences such as the contingent valuation method”; Hanley et al. [ 23 ] also outline choice experiments and analyze its roots in Lancaster’s characteristics theory of value; while the paper by Lancaster [ 38 ] is another relevant work in this group.

4. Discussion

Environmental valuation is intrinsically difficult because realistic environmental valuation situations are rarely observed, and singularities in environmental assets impede a uniform treatment of those values outlined by the Total Economic Value. Notwithstanding the difficulties, a plethora of papers have been published during the last decades.

As a result of this research it can be concluded that revealed preferences methodologies are surpassed by works focused on stated preference methods for the analyzed period as a whole. The research discloses the relevance of stated preference methods over revealed preferences methods, with a clear dominance of choice experiment over any other environmental valuation method, as predicted by Adamowicz [ 14 ]. The complexity of the choice experiment method has resulted in new challenges and research lines for academics. Choosing and implementing experimental designs, interpreting standard and more advanced random utility models, and estimating measures of willingness-to-pay are some of the issues covered by researchers [ 39 ].

Differences on the environmental valuation have been also revealed by the co-citation analysis, which reports different clusters by considering the methods used in the environmental valuation process. Despite its past influence, none of the travel cost and hedonic price methods is in the 10 most popular methods of environmental valuation, according to the keywords in the dataset used. In addition, the leading Journals in the publication of environmental valuation papers are ranked in prominent positions by WoS in their corresponding categories. The paper also distinguish two groups of authors according to the time they have published on the topic. The first group initiates the growth of the area in the mid-1990s, while the second group concentrates its impact mainly from 2010.

The abovementioned differences in the use of the environmental valuation methods do not imply that one method is unequivocally better or worse than another since its appropriateness depends on a particular situation. In other words, no single method is suitable in all valuation scenarios. Rather, the choice of the valuation method is context-specific. Revealed preference methods can be prioritized when budget and time are constrained. Stated preference methods require a complex questionnaire development and data analysis, which translates into an additional need of resources (both money and time). Conversely, revealed preference methods can only capture use values, while stated preference methods can estimate both use and non-use values. In addition, using multiple methodologies can be appropriate in some situations. For example, the combination of revealed and stated preference methods can improve benefit estimation of a single component [ 35 ]. This approach can be useful when a revealed preference method is utilized as the main valuation instrument, but some environmental values are more accurately estimated by using another method and the result is aggregated. In this case, the researcher must be careful to avoid double counting if the components of value captured by the different methods overlap [ 40 ].

5. Conclusions

From the evolution of environmental valuation publications in the last 30 years, we can assert that the discipline has been consolidated. Papers related to choice experiments have dominated academic production in the last decade. In the current stage of environmental valuation researchers will have to cope with new challenges and emerging trends. As in other research areas, the increasing ability to collect enormous amounts of data facilitates the creation of the available massive databases, which can be used to take environmental valuation methodologies to the next stage in their evolution by incorporating machine learning techniques in the valuation process. However, this evolution should not be restricted to new applications of the well-known valuation methods only. Researchers must develop new approaches to deal with new elements in the valuation process. We expect that climate change, as one of the defining challenges of the 21st century, will attract most attention from researchers to propose new approaches in environmental valuation [ 41 , 42 ]. As knowledge and perception are subjective, the intangible aspects must be explicitly considered in the new valuation methods [ 13 ]. In this regard, we may conclude that the future path of environmental valuation is not necessarily related to new methodologies, but to the inheritance and assimilation of consolidated techniques commonly used in other scientific areas.

Acknowledgments

We would like to thank three anonymous referees for their constructive comments and suggestions that substantially improved this article.

Abbreviations

The following abbreviations are used in this manuscript:

Author Contributions

Conceptualization, F.G. and P.T.; methodology, F.G. and P.T.; software, F.G.; validation, P.T.; formal analysis, F.G.; investigation, F.G. and P.T.; resources, P.T.; data curation, F.G.; writing—original draft preparation, F.G.; writing–review and editing, P.T.; visualization, F.G. and P.T.; supervision, P.T. All authors have read and agree to the published version of the manuscript.

This research received no external funding.

Conflicts of Interest

The authors declare no conflict of interest.

Advertisement

Economic valuation of wildlife conservation

• Published: 10 March 2023
• Volume 69 , article number  32 , ( 2023 )

Cite this article

• Simone Martino   ORCID: orcid.org/0000-0002-4394-6475 1 , 2 &
• Jasper O. Kenter   ORCID: orcid.org/0000-0002-3612-086X 1 , 3 , 4  

5372 Accesses

2 Citations

2 Altmetric

Explore all metrics

This paper reviews concepts and methods for the economic valuation of nature in the context of wildlife conservation and questions them in light of alternative approaches based on deliberation. Economic valuations have been used to set priorities, consider opportunity costs, assess co-benefits of conservation, support the case for conservation in public awareness and advocacy, and drive novel schemes to change incentives. We discuss the foundational principles of mainstream economic valuation in terms of its assumptions about values, markets, and human behaviour; propose a list of valuation studies in relation to wildlife protection; and explain the methods used. We then review critiques of these approaches focusing on the narrow way in which economics conceives of values, and institutional, power, and equity concerns. Finally, we complement conventional approaches commonly used for wildlife valuation with two forms of deliberative valuation: deliberated preferences and deliberative democratic monetary valuation. These are discussed in terms of their potential to address the drawbacks of mainstream economics and to realise the potential of valuation in bridging conservation of nature for its own sake and its important contributions to human well-being.

Similar content being viewed by others

Tourists’ valuation of nature in protected areas: A systematic review

Milena Gross, Jasmine Pearson, … Berta Martín-López

Economics of wildlife management—an overview

Ing-Marie Gren, Tobias Häggmark-Svensson, … Marc Engelmann

Conserving Tanzania’s Wildlife: What is the Policy Problem?

Avoid common mistakes on your manuscript.

Introduction

The importance of the social sciences in shaping conservation and wildlife management and in understanding the societal implications of conservation is increasingly recognised (Millennium Ecosystem Assessment 2005 ; Braat and de Groot 2012 ; Redpath et al. 2013 ; Sandbrook et al. 2013 ; Bennett et al. 2016 ,  2017 ). As a public good, wildlife has been heavily regulated including through the designation of protected species and protected areas, customs controls for trade in protected species, and animal welfare legislation. While there have been successes, the reality of many conflicting other policy priorities and conflicting economic incentives combined with weak institutions, limited resourcing and enforcement, and de facto open access regimes have led to many governance and market failures (Pearce and Moran 1994 ; Child et al. 2012 ). Also, regulatory approaches are only effective for those species and areas that are protected. This means additional approaches are necessary to ensure that the value of biodiversity more broadly is integrated into private and policy decisions. Valuing wildlife has emerged to counter perverse incentives to land use and management strategies (or lack thereof) that deplete natural resources and erode biodiversity (Daily and Ehrlich 1992 ; Pimentel et al. 1999 ). Many conservationists now recognize that economics needs to be a part of the design and implementation of conservation policies to achieve effective conservation whilst minimising opportunity cost for communities (Shogren et al. 1999 ; Barua et al. 2013 ; Emerton 1999 ; Sementelli et al. 2008 ), to maximise co-benefits of conservation in terms of ecosystem services that benefit human well-being (Tallis et al. 2008 ), Footnote 1 and to raise awareness of the value of nature and inventorise wildlife as natural capital stocks (Costanza et al. 2014 ; Jones et al. 2016 ). The proposal of wildlife as a natural asset having an economic value as other manufactured capitals can provide new perspectives to reduce human-wildlife conflicts (Nguyen et al. 2022 ), address the compatibility between conservation and hunting (Casola et al. 2022 ; Zhou et al. 2021 ; Gascoigne et al. 2021 ; Chapagain and Poudyal  2020 ; Travers et al. 2019 ; Fischer et al. 2015 ; Nielsen et al. 2014 ), generate new funding streams and support innovative financial tools (Emerton 1999 ; OECD 2013 ) that can facilitate the correction of market and governance failures (Pearce and Moran 1994 ). It is evident from what was said above that the adoption of economics in nature management reflects the vision of conservation policy based mainly on the instrumental (anthropocentric) value of nature rooted in the Hellenic and Judeo-Christian tradition (Parks and Gowdy 2013 ). These considerations show how nature conservation is becoming as much about people and institutions as it is about ecosystems and biodiversity (Mascia et al. 2003 ).

In the last 40 years, valuation studies on wildlife protection based on hypothetical markets and the analysis of revealed preferences mainly for uses such as recreation and tourism have proliferated (see Table 1 ). In addition, meta-analyses of biodiversity valuations have been undertaken to suggest what elements might influence and better direct an efficient allocation of resources for biodiversity protection (Loomis and White 1996 ; Martin-Lopez et al. 2007 ; Richardson and Loomis 2009 ).

However, due to the failure of policy initiatives to curb the loss of biodiversity and increasing emphasis on understanding multiple values in platforms such as the Intergovernmental Science Policy Platform on Biodiversity and Ecosystem Services (IPBES), new paradigms based on a closer synergy between “nature and people” (Mace 2014 ; Garcia-Jimenez et al. 2021 ) are advocated, and bio-cultural, deliberative and shared value approaches to conservation emphasise the need for pluralistic and partnership-based ways of conserving nature (Kenter 2016b ; Gavin et al. 2018 ; Christie et al. 2019 ; Diaz et al. 2018 ). Related to this, many critiques have been raised with traditional economic approaches. These include concerns about the public ability to understand and value biodiversity and what components can be feasibly expressed in monetary units (Nunes et al. 2001 ; Christie et al. 2006 ); concerns around the commodification of nature (McCauley 2006 ; Gómez-Baggethun and Ruiz-Pérez 2011 ); justice concerns around the prioritisation of certain values over others (Matulis 2014 ), and philosophical and psychological questioning of the narrow individualistic and utilitarian value assumptions of conventional (neoclassical) economics (O’Neill et al. 2008 ; Kenter et al. 2015 ). The emerging debate has given rise to novel approaches to valuation that seek to integrate shared, social and cultural values into valuations and decision-making. Here, group deliberation to form and bring together plural values plays an important role.

In this paper, we review the use of economic valuation in the context of wildlife conservation, outlining key concepts and methods illustrated with case examples identified through a structured literature review. Then we briefly review key critiques and concerns, before considering how recent deliberative approaches have attempted to address some of these concerns. The aim of this paper is to (1) provide an overview of economic valuation in a succinct and understandable way to those engaged in conservation science; (2) demonstrate the diverse way by which valuation is used to support wildlife protection in different contexts and regions; and (3) highlight novel deliberative approaches in valuation that could help bridge the increasing focus on human benefits of conservation with the value of wildlife as a good in and for itself. While there are various existing reviews of environmental valuation and a small number of reviews of deliberative monetary valuation (Spash 2008a , b ; Bunse et al. 2015 ), this is the first to specifically focus on wildlife conservation, and to provide an empirical overview of both traditional and deliberative economic studies.

Economic concepts to valuing biodiversity and wildlife

In the absence of human judgments, neoclassical economics suggests that goods and services in themselves have no value (Parks and Gowdy 2013 ), with valuation being the process of assessing preferences that express judgements on the value of goods. Therefore, only human preferences are recognised as the source of economic value (Mitchell and Carson 1989 ). Values of things, expressed as preferences, are thought to reflect the benefit of one additional unit of a good, keeping the amount of all other goods constant. This “marginalistic” notion of values relates to the conception of continuous exchange and trade-offs of goods or services within the market to maximise preference satisfaction (Lawson 2013 ). Trade-offs of goods against money are known as willingness to pay (WTP), the preferred measure of the intensity of preferences. Its counterpart, willingness to accept (WTA) provides another indication of our preferences towards goods in terms of how much we would need to receive in lieu of relinquishing something of value to us. Through this monetary expression of preferences, the market is thought to “invisibly” achieve the most efficient distribution of goods for society, optimising the allocation of scarce resources per those most willing to pay for them. Several further important assumptions are made. Preferences are considered purely individual and self-regarding, and neoclassical economics does not take an interest in what motivations underlay preferences, and values are considered pre-formed and rational. Value to society is established by aggregating individual values, usually through the simple addition of WTP (Hockley 2014 ).

When values associated with goods and services arising from nature are not accounted for in markets and policy because they are not traded, the resulting allocation of goods is inefficient. This is a situation that economists call market failure , and this has been an important impetus for extending the application of economic valuation to the environment. While adhering to a moral norm that biodiversity is basically good is an understandable position for many conservationists (Meffe and Caroll 1994 ), from a neoclassical economic perspective of value, only the biodiversity that provides benefits to humans that are marginally more valuable than the costs of implementing protection policies (including opportunity costs) should be protected (Roughgarden 1995 ). If conservation comes at the expense of alternatives that have a greater economic value to society, opportunity costs occur and justification for such a position would thus need to be made on non-economic grounds.

Typical examples of market failure include subsidisation of farming and ranching that make these activities more profitable than they would otherwise be, at the cost of undervalued wildlife (Emerton 1997 ; Child et al. 2012 ). Reverting this market distortion is possible by introducing different incentives motivating landowners to consider wildlife as a valuable asset that, if protected, can generate higher profits, including benefits from hunting that would make it compatible with conservation (Zhou et al. 2021 ; Chapagain and Poudyal  2020 ; Travers et al. 2019 ; Fischer et al. 2015 ; Nielsen et al. 2014 ) than land use that negatively affects those assets (e.g. intensive farming). This approach can also be applied to marine and terrestrial protected areas (McNeely 1993 ; Emerton 1998 ; Child, et al. 2012 ) that traditionally have been supported by public money. Thus, from an economic perspective, under non-distorted or corrected markets, competition for products and services drives prices of different uses for ecosystems by recognising the relative benefits of conservation and consequently making it a more attractive option (Loomis 1993 , 2000 ).

Environmental economists classify values according to the heuristic of Total Economic Value (Turner et al. 2003 ), which is split into two main categories: use and non-use values (Arrow et al. 1993 ). Use value includes non-consumptive use value such as that generated by recreational experience (Carr and Mendelsohn 2003 ; Becker et al. 2005 ; Getzner 2015 ; Chapagain and Poudyal  2020 ; Frew et al. 2018 ; Gascoigne et al. 2021 ) and consumptive use value such as from hunting (Berman and Kofinas 2004 ; Bennett and Whitten 2003 ). In addition, wildlife provides option value , the nominal price that one is willing to pay to safeguard potential future use (Pearce and Turner 1989 ). Non-use values, also called passive use values, include value for the sake of knowing that others ( altruistic value ), future generations ( bequest value ), or non-human nature ( existence value ) benefit (Ojea and Loureiro 2007 ; Saayman 2014 ; Morse-Jones et al. 2012 ; Molina et al. 2019 ). Altruistic, bequest, and existence values are, in neoclassical economics, still considered as self-regarding and not truly altruistic because it is solely the satisfaction to the person of having their preferences maximised that is assumed to be of value (Kenter 2015 ; Kenter et al. 2015 ; Kenter et al. 2016a ).

Finally, it is helpful to note that natural scientists and economists tend to consider biodiversity from different perspectives. While ecologists consider biodiversity as a representation of the complexity of the ecological system (Farnsworth et al. 2012 , 2015 ), the most common conceptions used by economists refer to specific biological components of diversity (Pearce and Moran 1994 ) such as the number or richness of species, threatened or endangered species, genetic resources of value, and ecosystem functions and services provided by ecosystems and single species (Bartkowski et al. 2015 ; Garcia-Jimenez et al. 2021 ; Markandya et al. 2008 ; Margalida et al. 2012 , 2010 ; Morales-Reyes et al. 2015 ; Becker et al. 2005 ). Consequently, environmental valuation typically refers to monetisation of the preferences in relation to these common indicators of biodiversity rather than systemic concepts of resilience and ecosystem stability (Christie et al. 2006 ; Farnsworth et al. 2015 ).

Economic valuation methods and applications to wildlife conservation

Building on the concepts discussed in the previous section, environmental economic valuation aims to assess individuals’ preferences, either stated or revealed (Turner et al. 2003 ). Stated preferences approaches, primarily including the contingent valuation method (CVM) (Hanemann 1994 ) and choice experiments (CE) (Birol et al. 2006 ), elicit respondents’ preferences in hypothetical scenarios, where they are asked either for their WTP for conservation, or (less commonly) their WTA for the loss of wildlife. They are the only method that is considered to effectively deal with non-use values (Arrow et al. 1993 ). In CVM, people are asked to directly state their WTP or WTA in relation to a clearly described environmental or policy change (e.g. for the conservation of charismatic species such as rhino (Saayman and Saayman 2017 ), Iberian and Eurasian lynx (Bartczak and Meyerhoff 2013 ; Molina et al. 2019 ) or sea turtle (Cazabon-Mannette et al. 2017 )). In CE, respondents are asked to consider several alternatives, described on the basis of multiple attributes, typically with a monetary cost attached to them and with a status quo at no cost. In CE, WTP is assumed to reflect the marginal rate of substitution between non-monetary attributes and the cost attribute. Several studies use CE to explore attributes of different policy measures that can reduce trade-offs between conservation, recreation, and hunting. To make a few examples, Bach and Burton ( 2017 ) use a choice experiment to explore trade-offs between conservation measures and recreational aspects of wildlife interactions with dolphins, while Nguyen et al. ( 2022 ) consider the possibility to reduce wildlife-human conflicts and Nielsen et al. ( 2014 ) and Travers et al. ( 2019 ) the possibility to implement measures that limit illegal hunting and trade of bush meat. Conventionally, stated preference methods are conducted via questionnaires, but there is increasing interest in deliberative monetary valuation (DMV) approaches (see “Discussion and conclusions” section), where people state their values after group discussions (Bunse et al. 2015 ).

Revealed preferences approaches use observed market data on an ordinary commodity to infer preferences for marginal changes in the quality or quantity of an environmental good associated with the ordinary commodity (Bockstael and McConnell 1993 ). Two key approaches include the hedonic pricing method (HPM) and travel cost method (TCM). With HPM the value of the environmental goods (e.g. the value of charismatic wildlife) is measured as a marginal impact on the value of immobile assets (typically premises or land) as in Casola et al. ( 2022 ) who investigated the marginal impact on property price caused by proximity and adjacency to hunting areas. In the case of TCM, preferences are revealed from money spent on travelling and the number of visits to a certain area to derive the recreational experience (Perman et al. 2003 ) as a measure of welfare accompanied by the economic impacts of recreation as direct spend and job creation (Frew et al. 2018 ; Fischer et al. 2015 ; Gascoigne et al. 2021 ).

Other approaches that are not based on preferences are also used in environmental valuation, though they are much less commonly used in relation to wildlife conservation. Production function approaches were commonly used during the 1990s to address the impact that natural capital might have on marketable outputs. Here the environment is treated as an “input” to the economic activity and its value is equated, like any other input, with the impact it has on the productivity of a marketed output. For example, Barbier ( 2000 ) looked at the value of regulating services such as nursery and habitat functions of mangroves in supporting the fishing shrimp industry in south Thailand and Mexico. Finally, approaches based on cost, such as avoided damage-cost, and replacement cost can be used for wildlife valuation. In the avoided damage-cost method, the benefits that are valued are considered at least equal to avoided damages, for example where natural predators control pests. An example of avoided cost provided by the conservation of vultures is shown by Markandya et al. ( 2008 ) who monetized the avoided cost of illness in terms of reduced bubonic and rabies diseases caused by stray dogs in India if vultures were protected. In the replacement cost method, minimum values of environmental goods are equated with the cost of replacing them with human solutions. For example, the cost of hiring honey beehives can be considered a proxy for the value of the services provided by wild pollinators (Sandhu et al. 2008 , 2015 ). Taxes and fines have also been used in some cases as a proxy for wildlife values, including non-use values (Sterner 2009 ), but they may be significantly less than values elicited through stated preference approaches. For example, the highest fine mandated by the 1973 US Endangered Species Act is $50,000, much lower than $173,209 considered as efficient to address the optimal level of protection of threatened and endangered organisms (Eagle and Betters  1998 ).

To illustrate the use of stated and revealed preference methods, we conducted a literature search of the last 30 years of research Footnote 2 on valuation considering method, species, type of value, geographic area, and motivations behind the study. The review was not meant to be fully comprehensive, but illustrative of the types of studies that have been undertaken to value wildlife. Searches were conducted on the Web of Science platform (searching in the fields “Title”, “Abstract”, and “Keywords”) using “wildlife” and “economics”, “contingent valuation”, “choice experiment” “travel cost method”, “hedonic price”, “deliberative monetary valuation” as search strings. Studies on habitat conservation, biodiversity, and in general terms to ecosystem services without reference to animal species were not included. In total, 150 studies emerged as based on conventional, non-deliberative stated, and revealed preference methods. A sample of these studies, organised by species and aim, is provided in Table 1 . The full set of studies is proposed in the Supplementary material. In addition, fourteen studies that conducted deliberative monetary approaches directly or indirectly related to wildlife protection were discussed separately in Section " Discussion and conclusions ".

Fifty per cent of the sample reported by our search used CVM, 27% used CE, 19% used TCM, 2% used HPM and the remaining 2% are more articulated valuation-based studies addressing the analysis of cost as a proxy for benefits, implementation of cos-benefit analysis with consideration of opportunity costs, or combination of market values with non-market price analysis. We found a relatively homogeneous distribution of studies (nearly 35 per continent) across Europe, the Americas, and Asia. Africa showed 25 studies, Oceania 22. CVM studies were dominating 20–30 years ago, now CE is equally prominent.

Table 1 shows how the literature has mainly focussed on mammals in temperate habitats (Johansson et al. 2012 ; Han and Lee 2008 ) and on the charismatic species of tropical countries (Saayman and Saayman 2014 ; Kaffashi et al. 2015 ; Rathnayake 2016 ; Wang et al. 2018 ), while less cases have been recorded in the marine (Saayman 2014 ; Batel et al. 2014 ; Pires et al. 2016 ; Nuno et al. 2018 ) and freshwater environments (Gan et al. 2011 ; Hutt et al. 2013 ).

Common patterns arising from Table 1 show that early studies (1980s and 1990s) addressed WTP drivers of wildlife conservation, while more recent research (since the 2000s) focuses more broadly on biodiversity and management tools such as taxes to protect bottlenose dolphin (Batel et al. 2014 ), hunting licences for elk and deer (Fix et al. 2005 ), and entrance fees for managing elephant conservation (Rathnayake 2016 ). Single species valuation is mainly carried out by CVM for the conservation of griffon vulture (Becker et al. 2005 , 2009 ), dolphin (Wang et al. 2016 ), catfish (Hutt et al. 2013 ), red grouse (Hanley et al. 2010 ), deer (Hanley et al. 2003 ), duck (Whitten and Bennett  2002 ), Eurasian lynx (Bartczak and Meyerhoff 2013 ; White et al. 2015 ) and Iberian lynx (Molina et al. 2019 ). Additional analysis refers to non-use value preservation of the big five (Kaffashi et al. 2015 ) and the big seven (Saayman and Saayman 2014 ).

CEs are used to relate wildlife conservation to contextual factors such as forest and wetland habitats management (Bergmann et al. 2006 ; Dias and Belcher  2015 ; Petrolia et al. 2014 ), or projects impacting the ecosystems and landscape such as renewable energies (Emmanouilides & Sgouromalli  2013 ; Ku and Yoo 2010 ; Lundhede et al. 2015 ). Use values of wildlife such as recreation (Zhou et al. 2021 ; Frew et al. 2018 ), viewing, and gaming are also frequently assessed by TCM (Gascoigne et al. 2021 ; Bennett and Whitten 2003 ), while only four papers have been found using HPM, all in relation to wildlife and hunters’ policy, assessing values of hunting licences (Lundhede et al. 2015 ; Getzner 2015 ; Mensah and Elofsson 2017 ; Casola et al. 2022 ). Studies addressing the relationship between wildlife viewing and hunting have shown benefits to individuals and society far greater than its realised economic earnings (Chapagain and Poudyal  2020 ; Frew et al. 2018 ; Fischer et al. 2015 ; Gascoigne et al. 2021 ), providing an opportunity to inform the value of taxes to better reflect consumer surplus arising from the recreational experience (Navrud and Mungatana 1994 ; Kaffashi et al. 2015 ); generating insights on new policy attributes that my help reduce the impact of illegal hunting and bush meat trade (Nielsen et al. 2014 ; Travers et al. 2019 ); raising awareness on the possible win–win solutions between conservation, recreation, and hunting (Gascoigne et al. 2021 ; Fischer et al.  2015 ); and addressing conflicts, to mention one specific case, between human communities and elephants (Nguyen et al. 2022 ). In addition, examples from southern Africa dryland ecosystems suggest that wildlife conservation in private ranches is becoming more remunerative than modern livestock production (Jenkins 2011 ; ABSA 2015 ) under communities’ right to use and benefit from wildlife (Child et al. 2012 ).

The literature explored revealed that CVM or CE is not mere econometric exercises. Studies frequently are designed for management purposes such as the implementation of market tools: entrance fees (Barnes et al. 1999 ; Fix et al. 2005 ; Stanley 2005 ; Mmopelwa et al. 2007 ; Cerda  2011 ; Kaffashi et al. 2015 ; Rathnayake 2016 ); property/user rights (Horne and Petajisto 2003 ; Sutton et al. 2008 ; Harihar et al. 2015 ); taxes to induce recreationalist behavioural changes in marine protected areas (Batel et al. 2014 ) and for terrestrial non-game conservation (Dalrymple et al. 2012 ). In addition, it is frequent the analysis of transaction and opportunity costs of wildlife conservation for communities (Barua et al. 2013 ; Emerton 1999 ; Sementelli et al. 2008 ) along with social valuations (Delibes-Mateos et al. 2022 ) in environmental cost–benefit analysis of policy scenarios, for instance, related to the analysis of the reintroduction of the Eurasian lynx in the UK (White et al. 2015 ; Hawkins et al. 2020 ).

Aside from the management implication of valuation, a set of studies investigated the motivations driving WTP, focusing on the role of socio-economic factors (Teal and Loomis 2000 ; Travers et al. 2019 ; Fischer et al. 2015 ; Nielsen et al. 2014 ), the knowledge of species endangerment (Tisdell and Wilson 2004 ; Tisdell et al. 2005a ; Morse-Jones et al. 2012 ), the degree to which people show appreciation for species (Tisdell et al. 2005b , 2007 ) or ethical aspects affecting WTP (Martinez-Espineira 2006 ; Ojea and Loureiro  2007 ).

Finally, benefit transfer methods were applied to inform benefits from a number of prior study sites to other locations where new primary studies are not feasible (Spash and Vatn 2006 ). A meta-analysis carried out by Loomis and White ( 1996 ) concluded that CVM provides estimates sensitive to the size of the endangered species population. Other meta-analyses identified a number of factors that help predict WTP for biodiversity conservation, including preferences for certain phylogenies (mammals and birds are preferred over reptiles) and economic effect of species such as damage to agriculture, which appeared to be more important than IUCN red list classification (Martin-Lopez et al. 2007 ; Richardson and Loomis 2009 ).

Notwithstanding the benefits of protection, wildlife conservation can be opposed and considered costly in case of predation on livestock, destruction of crops, traffic collisions, and transmission of diseases to animals and humans (Gren et al. 2018 ). Increased land use for agricultural purposes, occurring mainly in developing countries, also raises this type of conflict (Madhusudan 2003 ; Ninan and Sathyapalan 2005 ). However, there is evidence as reported in Table 1 of the benefits of conservation in comparison with the cost of protection such as for Griffon vulture (Johansson et al. 2012 ; Garcia-Jimenez et al. 2021 ; Markandya et al. 2008 ; Morales-Reyes et al. 2015 ; Becker et al. 2005 ), elephant (Sutton et al. 2008 ), deer (Bowker et al. 2003 ), and duck (Bennett and Whitten 2003 ). The latter examples show that valuation is often not an independent exercise, but more typically part of Cost Benefit Analysis (CBA), a way of appraising a policy by assessing the anticipated net benefits and costs (or ratio of benefits to costs) arising from its implementation. In conservation, benefits are often non-marketed values that can be addressed only by applying the methods above illustrated (Shwiff et al. 2013 ). These values are usually aggregated over a particular time span by discounting future values to the present, and for each affected party at stake, as shown for the reintroduction of the Eurasian lynx in the UK (White et al. 2015 ; Hawkins et al. 2020 ). Notwithstanding critiques of methods of valuation and aggregation, as reported in the next section, the range of economic approaches in Table 1 demonstrates the potential for economics to inform conservation policy and practice. This is achieved by an understanding of the human benefits and costs and potential trade-offs of conservation, and how incentives can change human behaviour and land management in a way that can enhance both conservation and economic outcomes.

Critiques of economic valuation

While the economic valuation of the benefits that wildlife provides to human well-being has been advocated as a powerful approach to enhance public, landowner, business, and policy support for conservation, it has equally received fierce criticisms. These critiques range broadly (for overviews see Forster 1997 ; Ravenscroft 2010 ; Kallis et al. 2013 ; Parks and Gowdy 2013 ; Matulis 2014 ; Schröter et al. 2014 ), but many of these issues can be related to either the narrow neoclassical economic conceptions of human values or to institutional questions of power and justice.

Value plurality

A key critique of mainstream environmental valuation relates to the conception of the consumer as a “rational agent” characterised by independent and exogenous preferences. It is highly questionable whether all values and behaviour can be couched in terms of pre-formed, self-regarding individual preferences that are rationally traded-off (Bowles and Gintis 2000 ; Balaine et al. 2020 ; Isacs et al. 2021 ). In relation to the natural environment, peoples’ motivations are very diverse, including rights, duties, virtues, cultural beliefs, identities, and narratives that are hard to translate into measures of utility (Spash 2006 ; Ojea and Loureiro 2007 ; Cooper et al. 2016 ). For example, people might be willing to pay for something because they feel it is the right thing to do, rather than because it satisfies their individual preferences. Moreover, many shared values operate at the level of communities and cultures, rather than individuals (Irvine et al. 2016 ), while the utilitarian framework assumes preferences can be counted in terms of individuals and added up. Even at the individual level, people may have multiple types of values that are difficult to compare. Ecological economists have proposed that irreducible value conflict is unavoidable and that only practical judgment formulated by deliberation provides an appropriate means for rational comparison (Martinez-Alier et al. 1998 ; Isacs et al. 2021 ). Furthermore, people may find it hard to understand biodiversity and ecosystems (Nunes et al. 2001 ; Christie et al. 2006 ; Hanley et al. 2015 ), often do not have clearly pre-formed values (Kenter et al. 2016c ), and may themselves prefer values to be expressed through group deliberation, to shape decisions under societal perspectives rather than that of individual consumers (Sagoff 1998 ; Ward 1999 ; Shapansky et al. 2003 ; Martinez-Espineira 2007 ; Bunse et al. 2015 ; Lienhoop et al. 2015 ; Kenter et al. 2016b ; Orchard-Webb et al. 2016 ).

Epistemologically, neoclassical economic approaches do not usually recognise pluralistic values and tend to exclude subjective and qualitative material (Parks and Gowdy 2013 ; Isacs et al. 2021 ). Economic studies focus on what is valued and how much , with little attention given to the why people value particular elements of nature in particular places. Economic conceptions and approaches for many violate a sense of integrity of nature, where the sacred, sublime, and aesthetic are reduced to mere numbers, and intrinsic values and rights of non-human nature are reduced to preferences (Cooper et al. 2016 ). It has been argued that the description of the environment in terms of preferences, utility, and WTP, at the least fails to reflect the deeper and often shared meanings that places might hold (Owen et al. 2009 ; Daniel et al. 2012 ) and at the worst is in itself a political act of commodification and enclosure (McCauley 2006 ; O’Neill et al. 2008 ; Gomez-Baggethun and Ruiz-Perez 2011 ; Turnhout et al. 2013 ; Matulis 2014 ; Spash 2015 ). A key risk of commodification is that, where human ingenuity can replace ecosystem services, the ecosystems providing those services lose their intrinsic value (Kronenberg 2015 ). However, other authors argue that a degree of commodification is necessary to embed the protection of services in a market context (TEEB 2010 ; UK NEA 2011 ) and to facilitate long-term behavioural changes towards nature conservation (Burton and Schwartz 2013 ).

Power, justice, and institutional critiques

In relation to conservation, there are many different dimensions of value that are difficult to trade-off against each other, raising the questions of power and justice in economic valuation. Such critiques of neoclassical economics focus on concerns with the role of cost–benefit analysis (CBA) in informing public policies (Hockley 2014 ) and assumptions that social welfare is measured by the sum of preferences of independent individuals (Parks and Gowdy 2013 ). When aggregating values, an agreement is needed on how to cluster within dimensions (i.e. how much does each individual count), and across dimensions of valuation (i.e. how are different value criteria to be made commensurate). Typically, CBA assumes that different groups in society have the same weight, but this can exacerbate social disparities with the poorest gaining less than the richest (Pearce et al. 2006 ). Another challenge is about what constitutes values: dimensions of value could be a set of straightforward economic or financial benefits and costs (e.g. expected revenue, construction, and operational costs), but the livelihoods of people, the cultural impact of the project, and impacts on local biodiversity cannot be easily monetised and used in CBA. In conventional economic analysis, if the benefits outweigh the costs after the compensation of losers, the project would be ‘efficient’ and deliver a net value to society (regardless of whether these compensations actually take place). This assumes that, in principle, the ecological, social, and cultural dimensions of value can be monetised and compensated fully and justly. However, unless all parties can agree on how different dimensions should be traded-off against each other, it is not possible to come to any single conclusion. Some economists themselves have argued that for this reason, economic valuation and appraisal have only limited use in complex and contested situations (Hockley 2014 ). People often resist attempts for their values to be converted into monetary amounts, and as a result, the use of economic approaches can increase conflict rather than resolve it when people feel that their other values are not taken into consideration (O’Brien 2003 ; Martino et al. 2019 , 2022 ). In more complex situations, collective choices taken in the interests of the community rather than economic transactions may dominate decisions and reduce the scope for economic valuation and CBA (Ostrom 1990 ).

To overcome these issues, frameworks of resource allocation and appraisal need to move beyond questions of preference utilitarian-based economic efficiency alone; for example, Turner ( 2016 ) proposes a ‘triple balance sheet’ framework to take account of shared and plural values, equity and justice, in those complex circumstances where the standard CBA is not considered adequate. This framework provides a basis for individuals and groups to work towards agreements from initially contesting positions (Bromley 2004 ), overcoming institutionalised economic analyses that are in most cases non-participatory (Christie et al. 2012 ) and limit their perceived democratic legitimacy (Hockley 2014 ). Even when CBA is appropriate, moving towards a more equitable CBA is now recognised as essential (Pearce et al. 2006 ; Turner 2007 , 2016 ) and examples are emerging, for instance in the context of climate change (van der Bergh and Botzen 2014 ; Nordhaus 2017 ).

Finally, it is important to note that there is confusion about the normative aim of valuation. The ultimate ecological goal of sustaining resilient ecosystems is not necessarily aligned with valuing biodiversity (Farnsworth et al. 2012 , 2015 ; Bartkowski et al. 2015 ). Although many environmental economists, and increasingly conservationists, see valuation as a means to an end, whereby the end is conservation (Spash and Aslaksen 2015 ), this end is different from economists’ goal of informing implications for human welfare, albeit defined within the narrow terms of utility and efficiency. While many economists seem to think the two can be harmonised (Kenter 2016b ), economic analysis may not favour conservation action. For example, van Beukering et al. ( 2014 ) valued the benefits and costs of a project to eradicate invasive species from a conservation area in Montserrat through a choice experiment approach but found that people’s preferences were lined up against this. This resonates with fundamental questions about whether the common good, as an objective of public policy, can really be equated with aggregate benefits to individuals (Sagoff 1998 ), and the degree to which individual preferences can account for a long-term sustainability perspective, also where understanding of ecological processes is limited to experts. Thus, in conventional economic approaches, when people’s utilitarian preferences do not align with long-term sustainability, we cannot expect approaches based on such preferences to lead to sustainable outcomes (Norgaard 2010 ; Everard et al. 2016 ).

• Deliberative monetary valuation

To address the issues above described, several different ways forward are being pursued. These include (1) the development of frameworks of environmental values that move beyond individualism and utilitarianism (Spash 2008b ; Chan et al. 2012 , 2016 ; Kenter et al. 2015 , 2019 ; Scholte et al. 2015 ; Pascual et al. 2017 ; O’Connor and Kenter 2019 ; Kenter and O’Connor 2021 ); (2) non-monetary valuation approaches, as an alternative to economic ones (Raymond et al. 2014 ; Ranger et al. 2016 ), to generate multiple evidence bases (Tengö et al. 2014 ; Hattam et al. 2015 ), or integrated with monetary valuation (Jacobs et al. 2016 ; Kenter 2016c ; Kenter et al. 2016b ); (3) development of deliberative valuation institutions (MacMillan et al. 2002 , 2006 ; Spash 2007 , 2008a , b ; Lo and Spash 2012 ; Kenter et al. 2014 , 2016a , c ; Bunse et al. 2015 ; Orchard-Webb et al. 2016 ; Kenter 2017 ; Bartkowski and Lienhoop 2018 ; Schaafsma et al. 2018 ).

Considering the focus of the paper on economic valuation, the interest here shifts to an emerging set of methodologies termed deliberative monetary valuation (DMV), a term first employed by Spash ( 2007 ). Deliberative approaches are a way to clearly form and express preferences (e.g. Alvarez Farizo et al. 2007 ; Lliso et al. 2020 ), providing an opportunity to engage with a broader array of motivations and moral stances than utilitarian individual preferences alone (e.g. Kenter et al. 2011 ). Also, most biodiversity is set in a non-western context, where conventional survey-based approaches may be inappropriate or ineffective because of cultural differences and lower levels of literacy that can be addressed by linking participation and valuation through deliberation (Christie et al. 2012 ). DMV approaches integrate participation, reflection, discussion, and social learning into the monetary valuation of environmental and other public goods, or budgetary decisions relating to the provisioning of such goods. In DMV, small groups of participants explore values and preferences for different policy options through reasoned discourse. Values may be expressed (Table 2 ) distinguishing across two dimensions (Spash 2007 ): the value provider (individual vs. group) and scale (individual vs. societal). This schematic thus suggests four main types of monetary value indicator: (i) a deliberated individual WTP, (ii) a ‘fair price’, and a deliberated social WTP (i.e. value to society) determined by either (iii) individuals or (iv) the group. Most DMV studies have used individual WTP (Bunse et al. 2015 ), including in wildlife conservation (Table 3 ). In contrast, social WTP constitutes a pre-aggregated value to society, established through consensus or voting, on how much participants think society should spend on one thing over another, by deciding how much of a budget should be allocated to the provisioning of different public goods. In a fair price payment valuation, participants state what they think both others and themselves should pay, and thus also brings in more of a moral context than conventional elicitation of individual WTP. Moreover, the deliberative forum provided by DMV provides an opportunity to link monetary valuation to non-monetary methods where participants can express and relate values more broadly. For example, the UK National Ecosystem Assessment, whilst establishing the value of protecting wildlife and habitats in marine protected areas, linked monetary valuation to storytelling and wellbeing indicators to consider experiences, identities, and capabilities arising from the places people visited, as well as a values compass to reflect on the broader principles and life goals that MPA management should be aligned with (Bryce et al. 2016 ; Kenter et al. 2016b ).

Kenter ( 2017 ) divides DMV studies into two main types, though some have features of both: deliberated preferences approaches, where deliberation is integrated into stated preferences, and deliberative democratic monetary valuation (DDMV), where assumptions of neoclassical economics are relaxed and the focus is to deliberate directly on the common good. Deliberated preferences studies focus on providing research participants time to discuss and think about their preferences, and help participants become more familiar with the environmental goods they are being asked to value (Christie et al. 2006 ). This concern was the main motivation for some of the first empirical DMV papers, where DMV was conceptualised as a “market stall” facilitating participants to become familiar with the goods they had to value, as first applied by Macmillan et al. ( 2002 ) to study the value of wild goose conservation in Scotland.

Table 3 provides an overview of DMV studies relevant to wildlife conservation, most of which constitute deliberated preference approaches, linking CVM or more occasionally CE to deliberation. We found only seven cases of wildlife valuation via DMV (McMillan et al. 2002 ; McMillan et al. 2006 ; Philips and McMillan 2005 ; Lienhoop and Fischer 2009 ; Watzold et al. 2008 ; Kenter 2016c ; Kenter et al. 2016b ), although other studies are related to wildlife in the broader context of conservation in agriculture, marine, forest, river, and wilderness environments (Lienhoop and McMillan 2007 ; Zsabo 2011 ; Vargas and Diaz 2017 ; Vargas et al. 2017 ; Kenter et al. 2011 ; Lliso et al. 2020 ; Orchard-Webb et al. 2016 ). Deliberated preferences are focusing on the conservation of broader biodiversity and ecosystem services and the scope for preferences elicitation in the context of complex decisions affecting multiple environmental attributes (also see Bunse et al. 2015 ; Schaafsma et al. 2018 ; Lliso et al. 2020 ). Another context where deliberated preferences are promising is in relation to unknown underwater wildlife and habitats, where the absence of preformed preferences is even more of an issue (Spash 2002 ; Hanley et al. 2015 ; Jobstvogt et al. 2014a , b ). A DMV study with scuba divers and sea anglers demonstrated that even these expert participants, who were much more familiar with marine environments than the general public, lacked well-developed preferences, but they were able to develop habitat-specific preferences through group discussion (Kenter et al. 2016a ).

To address the critiques in relation to value plurality, institutions, equity, and power discussed in the previous section, DDMV approaches make a more radical departure from conventional economic assumptions. In DDMV participants consider the benefits and costs of different policy options alongside non-instrumental concerns, including social norms, rights, and duties, virtues such as fairness or responsibility, and narratives—stories that can implicitly and explicitly express values (Kenter 2017 ). While this process has been termed preference moralisation (Lo and Spash 2012 ), the values involved are more than just moral values because they address broader conceptions of what is important in life, what Kenter et al. ( 2015 ) and Raymond and Kenter ( 2016 ) term transcendental values, relating to shared communal, cultural and societal values and also to the relations between environment and culture. These values are often latent, emphasising the need for explicit consideration in deliberation. Evaluation in DDMV takes place through communicative rather than instrumental rationality, where the common good is conceived of as ultimately a question of communication and deliberation to find common agreement (Orchard-Webb et al. 2016 ).

There has been one DDMV study referring to wildlife, where participants in deliberation on hypothetical policy options for strategic local development around fisheries and the broader coastal environment established a social WTP (Orchard-Webb et al. 2016 ). This demonstrated that participants can indeed set their individual utility aside to negotiate social WTP for policy options that reflect a range of value types and concerns. The key ground on which to assess whether such valuations are rational is whether all salient perspectives and interests have been included, and outcomes are not distorted by power relationships (Howarth and Wilson 2006 ). The aim is to create a democratic platform for evaluating options across different types of ethical and practical stances and for integrating intrinsic values of wildlife into the valuation process because the final value indicators are not conceived of as representing the sum of individual utilities but as an expression of shared values and social priorities.

Thus, an important concern with DDMV and DMV generally is who sits at the table and how can their inclusive, noncoercive participation be ensured (Kenter 2017 ; Schaafsma et al. 2018 ; Zimmerman et al. 2021 ). Fortunately, DMV can build here on well-established traditions of broader deliberative and participatory research and practice in e.g. political science, development studies, and stakeholder participation in environmental management (e.g. Chambers 1997 ; Jordan 2014 ; Devente et al. 2016 ), and guidance for best practice is starting to emerge (Kenter et al. 2016a , b , c ; Schaafsma et al. 2018 ).

Adapted with modifications from Kenter ( 2017 ), originally adapted with modifications from Spash ( 2007 , 2008a , b ).

Discussion and conclusions

This paper has discussed the role of economic valuation in the context of wildlife conservation by proposing a broad range of studies from several geographic regions and for different environmental domains and species. The proposed review can provide to the community of researchers and practitioners operating in the field of wildlife conservation an important instrument for the dissemination and synthesis of the existing literature on economic valuation with implications for policy and regulating mechanisms and tools necessary for the protection of wildlife. In addition, this paper, for the first time, to the best knowledge of the authors, suggests looking at an economic valuation under a different lens by proposing key critiques of traditional mainstream valuation methods and considering how recent deliberative approaches are able to address some of the concerns raised.

The first part of this review has dealt with the role of classical economics methods for measuring the welfare of wildlife conservation and suggested how economic valuation of wildlife can help address market failures that lead to the depletion of finite natural resources and the decline of biodiversity (Daily and Ehrlich 1992 ; Pimentel et al. 1999 ) through the elicitation of environmental values driven by people’s behaviour. These methods have been used, among others, to set priorities and innovative policy attributes that show people’s motivation for conservation against resources depredation (Gascoigne et al. 2021 ; Delibes-Mateos et al. 2022 ; Nielsen et al. 2014 ; Travers et al. 2019 ; Zhou et al. 2021 ), to promote awareness of the value of nature and wildlife as natural capital stocks (Costanza et al. 2014 ; Jones et al. 2016 ), to generate new funding streams and financial tools (Emerton 1999 ; OECD 2013 ) and to set taxes and fees for promoting behavioural change (Emerton 1999 ; Sterner 2009 ; Batel et al. 2014 ; Batel et al. 2014 ; Dalrymple et al. 2012 ) in order to correct market failures (Pearce and Moran 1994 ).

Many studies have been promoted to support the case for conservation in public awareness and advocacy, facilitating where possible the setting of policy measures that may induce the coexistence between conservation, recreation, and hunting. Recent examples from around the world have used valuation approaches for addressing conservation-hunting duality in species like duck (Bettett and Whitten 2003 ), swan (Frew et al. 2018 ), elk (Chapagain and Poudyal  2020 ), lynx (Molinaet al. 2019 ; Bartczak and Meyerhoff 2013 ), grey wolf (van Eeden et al.  2021 ) and brown bear (Richardson and Lewis 2022 ) to mention a few, and reduction of wildlife-human conflicts addressing the balance between benefits and costs of reintroducing the Eurasian lynx in the UK (White et al. 2015 ; Hawkins et al. 2020 ) and the Iberian lynx in Spain (Delibes-Mateos et al. 2022 ).

Economic analysis has been used to show that wildlife is an asset that, if protected, can generate substantial economic benefits for those who have the right to own or use it (Loomis 1993 , 2000 ). It is not surprising how the recreational aspects of wildlife have promoted valuation studies on mammals in temperate habitats and charismatic tropical species, with less attention to marine and aquatic species and habitats. However, numerous economic reasons exist for a more comprehensive preservation strategy, including regulating services and diverse cultural aspects, that considers habitat-wildlife interactions more broadly (de Groot et al. 2002 ; MEA 2005 ; TEEB 2010 ; UK NEA 2011 ; IPBES 2019 ). Recent studies on the ecosystem services provided by vultures show the high cost of policies failing to address the reduction of these wild birds in terms of avoided sanitary benefits due to the spread of dogs in the global south (Markandya et al. 2008 ; Margalida et al. 2010 , 2012 ), as well as the loss of regulating and supporting services such as the nutrient cycle by processing animal carcasses that if treated by current solid waste processes would contribute to generating pollution and carbon dioxide emissions (Morales-Reyes et al. 2015 ). Studies addressing the regulating services provided by vultures are so far limited to a few Asian countries and in Israel (Becker et al. 2005 , 2009 ), but their diffusion can facilitate the adoption of cost-effective conservation policies along with the implementation of more recent findings on the cultural ecosystem services provided by these wild birds whose protection can sustain the local economy of communities, such as those living in the Pyrenees (Spain), by the enhancement of avian scavenger-based tourism (Garcia-Jimenez et al. 2021 ).

Although many of the proposed studies were designed to address policies for conservation, there are examples where valuation does not result in economic preferences for conservation (e.g. van Beukering et al. 2014 ). This may happen when people are not familiar enough with ecosystems to express robust values for their protection (McMillan et al. 2006 ). Moreover, the conception of values in mainstream economic valuation is narrow, does not consider shared values (Kenter et al. 2015 ), fails to reflect the deeper meanings that places might hold (Daniel et al. 2012 ), and for many violates a sense of integrity of nature (McCauley 2006 ; O’Neill et al. 2008 ; Kronenberg 2015 ; Cooper et al. 2016 ).

Attempts to trade-off between ecological, social, and cultural dimensions of the value of nature by assuming these values are commensurable can increase conflicts rather than resolve them. Moreover, decisions based on CBA, even when based on comprehensive transaction and opportunity cost of wildlife conservation (Emerton 1999 ; Sementelli et al. 2008 ; Barua et al. 2013 ) to define value for money of policy scenarios, as for the reintroduction of species like the Eurasian lynx (White et al. 2015 ; Hawkins et al. 2020 ), can in some circumstances implicitly supports the rich and powerful, because in monetary terms they will have the largest benefits (Hockley 2014 ; Turner 2016 ). Nonetheless, understanding social perspectives on the allocation of scarce resources to conservation efforts remains important. There is substantial opportunity to consider more pluralistic values within economic valuation through DMV and better integration with non-monetary approaches. For example, the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES 2019 , 2022 ), has been exploring ways to integrate the knowledge, values, and rights of local and Indigenous Peoples into conservation initiatives. The latter is an expression of a bio-cultural approach to conservation that emphasises the role of those who should be involved in the conservation process, and the need for pluralistic and partnership-based approaches to conservation (Gavin et al. 2018 ).

Deliberated preferences methodologies can support participants to think through and discuss values with each other, providing opportunities to enhance understanding of biodiversity and ecosystems through social learning. By moving from individual attitudes (Kahneman 2011 ) to preferences constructed by deliberation, as promoted by environmental psychology, human ecology, and ecological economics (Mascia et al.  2003 ; Saunders et al. 2006 ), there is potential for a more robust assessment of values for wildlife conservation.

The implementation of DMV may in fact change completely the perception of values from a finite to an inestimable asset (Kenter et al. 2011 ), limiting the use of CBA. When goods become priceless, CBA is no more of utility for decision making thus advocating the use of pluralistic and partnership-based approaches to conservation such as the Triple balance approach (Turner 2016 ). This framework suggests that only in simple and non-contested contexts, a single balance sheet consisting of a modified version of CBA is adequate. When there is increased complexity and contestation, two further sheets can be added that address social impact assessment and broader shared and cultural values. Values are aggregated and value conflicts are negotiated through deliberation. In addition, DMV studies have contributed to better exploring unrecognised preferences for wildlife conservation compared to more acknowledged projects such as renewable energy production (McMillan et al. 2006 ) and to estimating motivation for entering payment for ecosystem services schemes for the protection of nature (Lliso et al. 2020 ).

DDMV goes further as a platform to deliberate directly on the common good through discussion of social willingness to pay, providing new democratic spaces based on deliberative democracy (Irvine et al. 2016 ; Kenter 2016b ; Kenter et al. 2016a ). However, there are yet few deliberated preference studies, and a very limited number of DDMV studies informing practical conservation on the ground. Future research is necessary to demonstrate whether different DMV approaches can meet the practical demands of conservation practitioners in terms of planning, priority-setting, advocacy, and providing effective incentives. However, some of the signs are promising. While some have argued that strategies that recognize and work within the boundaries of existing values are more productive than trying to change values (Manfredo et al. 2016 , 2017 ), deliberative interventions can and do change values towards a more sustainability-aligned perspective, both at the contextual level (specific values reflecting opinions of the importance of something) and transcendental level (broad overarching guiding principles and life goals) (Kenter et al. 2011 , 2016b ; Kenter 2016c ; Raymond and Kenter 2016 ). However, whether this happens will depend on whether deliberation specifically targets transcendental values and more broadly is able to make previously implicit values explicit (Kenter et al. 2016c ).

DMV and DDMV particularly provide more of an opportunity to consider intrinsic values of nature alongside instrumental ones, but so far there has not been researched to explicitly and fully realise this, providing a potentially powerful opportunity to address an important critique in relation to economic valuation. In addition, the integration of participatory approaches may provide a more effective way to integrate values into policy and practice on the ground. As such, the future of environmental valuation may be in a closer integration between economic and broader social science approaches, which will require a loosening of some of the epistemic, ethical, value, and rationality assumptions that neoclassical environmental economists have upheld so far. However, this requires important advancement in terms of upscaling deliberative valuations, further development of best practices for deliberative interventions to support social learning and formation of shared values, and application across a more diverse range of conservation scenarios and institutional venues (e.g. payments for ecosystem services, community conservation agreements, protected areas establishment/management), and developing culturally appropriate methodological approaches for application in diverse contexts in the global south. If this can be achieved, by opening to broader perspectives of how people value and relate to nature, valuation could yet become an important bridging instrument between conserving nature for its own sake and for its important contributions to human well-being.

Data Availability

All data used were provided and made available in the manuscript. They were extracted by the bibliography and reported in all the tables included in the manuscript and in the additional references reported in the supplementary material.

For instance, some benefits provided by vulture have been quantified in biophysical terms and monetary units pointing on the capacity of these wild birds to provide regulating (avoidance of carbon emission) and cultural ecosystem services (e.g., recreation) as proposed by Garcia-Jimenez et al. ( 2021 ), Morales-Reyes et al. ( 2015 ) and Becker et al. ( 2005 ).

The initial search at the time of submission (May 2022) considered literature searches from 1990 until 2019 reflecting delay in the submission of the paper caused by the COVID-19 pandemic. Under the revision if the paper, literature has been extended using the same search criteria to consider papers published until December 2022.

ABSA (2015) Game Ranching Profitability in South Africa, ABSA & Barclays

Ali AHM, Afandi SHM, Emmy PJ, Shuib A, Ramachandran S, Samdin Z (2018) Assessment of non-consumptive wildlife-oriented tourism in Sukau, Sabah using travel cost method. Int J Bus Soc 19(1):47–55

Google Scholar  

Alvarez Farizo B, Hanley N, Barberán R, Lázaro A (2007) Choice modeling at the market stall: individual versus collective interest in environmental valuation. Ecol Econ 60:743–751

Article   Google Scholar  

Arrow K, Solow R, Portney PR, Leamer E, Radner R, Schuman H (1993) Report of the NOAA Panel on Contingent Valuation. Resources for the Future, Washington, DC, p 38

Bach L, Burton M (2017) Proximity and animal welfare in the context of tourist interactions with habituated dolphins. J Sustain Tour 25(2):181–197

Balaine L, Gallai N, Del Corso JP, Kephaliacos C (2020) Trading off environmental goods for compensations: insights from traditional and deliberative valuation methods in the Ecuadorian Amazon. Ecosyst Serv 43:101110. https://doi.org/10.1016/j.ecoser.2020.101110

Bandara R, Tisdell C (2004) The net benefit of saving the Asian elephant: a policy and contingent valuation study. Ecol Econ 48(1):93–107

Barbier E (2000) Valuing the environment as input: applications to mangrove- fishery linkages. Ecol Econ 35:47–61

Barnes JI, Schier C, van Rooy G (1999) Tourists’ willingness to pay for wildlife viewing and wildlife conservation in Namibia. S Afr J Wildl 29(4):101–111

Bartczak A, Meyerhoff J (2013) Valuing the chances of survival of two distinct Eurasian lynx populations in Poland e Do people want to keep the doors open? J Env Manage 129:73–80

Bartkowski B, Lienhoop N, Hansjürgens B (2015) Capturing the complexity of biodiversity: a critical review of economic valuation studies of biological diversity. Ecol Econ 113:1–14

Bartkowski B, Lienhoop N (2018) Beyond rationality, towards reasonableness: enriching the theoretical foundation of deliberative monetary valuation. Ecol Econ 143:97–104. https://doi.org/10.1016/j.ecolecon.2017.07.015

Barua M, Bhagwat SA, Jadhav S (2013) The hidden dimensions of human–wildlife conflict: Health impacts, opportunity and transaction costs. Biol Conserv 157:309–316

Batel A, Basta J, Mackelworth P (2014) Valuing visitor willingness to pay for marine conservation. The case of the proposed Cres-Losinj Marine Protected Area. Croatia Ocean Coast Manag 95:72–80

Becker N, Choresh Y, Bahat O, Inbar M (2009) Economic analysis of feeding stations as a means to preserve an endangered species: the case of Griffon Vulture (Gyps fulvus) in Israel. J Nat Conserv 17(4):199–211

Becker N, Inbar M, Bahat O, Choresh Y, Ben-Noon G, Yaffe O (2005) Estimating the economic value of viewing griffon vultures Gyps fulvus: a Travel Cost Model Study at Gamla Nature Reserve. Israel ORYX 39(4):429–434

Bennett NJ, Roth R, Klain SC, Chan KMA, Clark DA, Cullman G, Epstein G, Nelson MP, Stedman R, Teel TL, Thomas R, Wyborn C, Curran D, Greenberg A, Sandlos J, Verıssimo D (2016) Mainstreaming the social sciences in conservation. Conserv Biol 31(1):56–66

Article   PubMed   Google Scholar  

Bennett NJ, Roth R, Klain SC, Chan K, Christie P, Clark DA, Cullman C, Curran D, Durbin TJ, Epstein G, Greenberg A, Nelson MP, Sandlos J, Stedman R, Teel TL, Thomas R, Veríssimo D, Wyborn C (2017) Conservation social science: understanding and integrating human dimensions to improve conservation. Biol Conserv 205:93–108

Bennett J, Whitten S (2003) Duck hunting and wetland conservation: compromise or synergy? Can J Agric Econ 51(2):161–173

Bergmann A, Hanley M, Wright R (2006) Valuing the attributes of renewable energy investments. Energy Policy 34(9):1004–1014

Berman M, Kofinas G (2004) Hunting for models: grounded and rational choice approaches to analysing climate effects on subsistence hunting in an Arctic community. Ecol Econ 49(1):31–46

Birol E, Karousakis K, Koundour P (2006) Using a choice experiment to account for preference heterogeneity in wetland attributes: the case of Cheimaditida wetland in Greece. Ecol Econ 60(1):145–156

Bockstael N, McConnell K (1993) Public goods as characteristic of non-market commodities. Econ J 103:1244–1257

Bond CA, Cullen KG, Larson DM (2009) Joint estimation of discount rates and willingness to pay for public goods. Ecol Econ 68(11):2751–2759

Bosetti V, Pearce D (2003) A study of environmental conflict: the economic value of Grey Seals in southwest England. Biodivers Conserv 12(12):2361–2392

Bowker JM, Newman DH, Warren RJ, Henderson DW (2003) Estimating the economic value of lethal versus nonlethal deer control in suburban communities. Soc Nat Resour 16(2):143–158

Bowles S, Gintis H (2000) Walrasian economics in retrospect. Q J Econ 115:1411–1439

Boxall PC, Macnab B (2000) Exploring the preferences of wildlife recreationists for features of boreal forest management: a choice experiment approach. Can J for Res 30(12):1931–1941

Braat LC, de Groot R (2012) The ecosystem services agenda: bridging the worlds of natural science and economics, conservation and development, and public and private policy. Ecosyst Serv 1:4–15

Brock M, Perino G, Sugden R (2017) The warden attitude: an investigation of the value of interaction with everyday wildlife. Environ Resour Econ 67(1):127–155

Bromley DW (2004) Reconsidering environmental policy: prescriptive consequentialism and volitional pragmatism. Environ Resour Econ 28:73–99

Bryce R, Irvine KN, Church A, Fish R, Ranger S, Kenter JO (2016) Subjective well-being indicators for large-scale assessment of cultural ecosystem services. Ecosyst Serv 21:258–269

Bunse L, Rendon O, Luque S (2015) What can deliberative approaches bring to the monetary valuation of ecosystem services? A literature review. Ecosyst Serv 14:88–97

Burton RJF, Schwartz G (2013) Result oriented agri-environmental scheme and their potential for promoting behavioural changes. Land Use Policy 30(1):628–641

Callaghan CT, Slater M, Major RE, Morrison M, Martin JM, Kingsford RT (2018) Travelling birds generate ecotravellers. The economic potential vagrant. Hum Dimens Wildl 23(1):71–82

Carr L, Mendelsohn R (2003) Valuing coral reefs: a travel cost analysis of the Great Barrier Reef. Ambio 32(5):353–357

Casola WR, Peterson MN, Wu Y, Sills EO, Pease BS, Pacifici K (2022) Measuring the value of public hunting land using a hedonic approach. Hum Dimens Wildl 27(4):343–359. https://doi.org/10.1080/10871209.2021.1953196

Cazabon-Mannette M, Schumann P, Hailey A, Horrocks J (2017) Estimating the non-market value of sea turtle in Tobago using stated preferences techniques. J Environ Manage 192:281–291

Cerda C (2011) Willingness to pay to protect environmental services: a case study with use and non-use values in central Chile. Interciencia 36(11):796–802

Cerda C, Losada T (2013) Assessing the value of species: a case study on the willingness to pay for species protection in Chile. Environ Monit Assess 185(12):10479–10493

Chambers R (1997) Whose Reality Counts? Putting the First Last. ITDG

Chan KMA, Balvanera P, Benessaiah K, Chapman M, Díaz S, Gómez-Baggethun E, Gould R, Hannahs N, Jax K, Klain S, Luck G, Martín-López B, Muraca B, Norton B, Ott O, Pascua U, Satterfield T, Tadaki M, Taggart J, Turn N (2016) Opinion: why protect nature? Rethinking values and the environment. PNAS 113:1462–1465

Article   CAS   PubMed   PubMed Central   Google Scholar  

Chan KMA, Satterfield T, Goldstein J (2012) Rethinking ecosystem services to better address and navigate cultural values. Ecol Econ 74:8–18

Chapagain BP, Poudyal NC (2020) Economic benefit of wildlife reintroduction: a case of elk hunting in Tennessee, USA. J Environ Manage 269-110808

Child BA, Musengezi J, Parent GD, Child GFT (2012) The economics and institutional economics of wildlife on private land in Africa. Pastoralism 2:18

Chriestie M, Martin-Lopez B, Church A, Siwicka E, Szymonczyk P (2019) Understanding the diversity of values of “Nature’s contributions to people”: insights from the IPBES assessment of Europe and central Asia. Sustain Sci 14:1267–1282

Christie M, Fazey I, Cooper R, Hyde T, Kenter JO (2012) An evaluation of monetary and non-monetary techniques for assessing the importance of biodiversity and ecosystem services to people in countries with developing economies. Ecol Econ 83:69–80

Christie M, Hanley N, Warren J, Murphy K, Wright R, Hyde T (2006) Valuing the diversity of biodiversity. Ecol Econ 58:304–317

Cooper N, Brady E, Steen H, Bryce R (2016) Aesthetic and spiritual values of ecosystems: recognising the ontological and axiological plurality of cultural ecosystem “services.” Ecosyst Serv 21:218–229

Cornicelli L, Fulton DC, Grund MD, Fieberg J (2011) Hunter perceptions and acceptance of alternative deer management regulations. Wildl Soc Bull 35(3):323–329

Costanza R, de Groot R, Sutton P, van der Ploeg S, Anderson SJ, Kubiszewski I, Farber S, Turner RK (2014) Changes in the global value of ecosystem services. Glob Environ Change 26:152–158

Curtis JA (2002) Ethics in wildlife management: What price? Environ Values 11(2):145–161

Daily GC, Ehrlich PR (1992) Population, sustainability, and Earth’s carrying capacity: A framework for estimating population sizes and lifestyles that could be sustained without undermining future generations. Bioscience 42(10):761–777

Dalerum F, Miranda M, Muniz C, Rodriguez P (2018) Effect of scarcity, aesthetics and ecology on wildlife auction prices of large African mammals. Ambio 47(1):78–85

Article   CAS   PubMed   Google Scholar  

Dalrymple CJ, Peterson MN, Cobb DT, Sills EO, Bondell HD, Dalrymple DJ (2012) Estimating Public Willingness to Fund Nongame Conservation through State Tax Initiatives. Wildl Soc Bull 36(3):483–491

Daniel TC, Muhar A, Arnberge A, Aznar O, Boyd JW, Chan K, Costanza R, Elmqvist T, Flint C, Gobste P, Grêt-Regamey A, Lave R, Muhar S, Penker M, Ribe R, Schauppenlehner T, Sikor T, Soloviy I et al (2012) Contributions of cultural services to the ecosystem services agenda. PNAS 109:8812–8819

Delibes-Mateos M, Glikman JA, Lafuente R, Villafuerte R, Garrido FE (2022) Support to Iberian lynx reintroduction and perceived impacts: assessments before and after reintroduction. Conserv Sci Pract 4:e605. https://doi.org/10.1111/csp2.605

de Groot RS, Wilson MA, Boumans RMJ (2002) A typology for the classification, description and valuation of ecosystem functions, goods and services. Ecol Econ 41:393–408

Devente J, Reed MS, Stringer LC, Valente S, Newig J (2016) How does the context and design of participatory decision making processes affect their outcomes? Evidence from sustainable land management in global drylands. Ecol Soc 21(2):24. https://doi.org/10.5751/ES-08053-210224

Dias V, Belcher K (2015) Value and provision of ecosystem services from prairie wetlands: a choice experiment approach. Ecosyst Serv 15:35–44

Díaz S, Pascual U, Stenseke M, Martín-López B, Watson RT, Molnár Z, Hill R, Chan KMA, Baste IA, Brauman KA, Polasky S, Church A, Lonsdale M, Larigauderie A, Leadley PW, Van Oudenhoven APE, van der Plaat F, Schröter M, Lavore S, Aumeeruddy-Thomas Y, Bukvareva E, Davies K, Demissew S, Erpul G, Failler P, Guerra CA, Hewitt CL, Keune H, Lindley S, Shirayama Y (2018) Assessing nature’s contributions to people. Science 359:270–272. https://doi.org/10.1126/science.aap8826

Eagle JG, Betters DR (1998) The endangered species act and economic values: a comparison of fines and contingent valuation studies. Ecol Econ 26:165–171

Emerton L (1997) The economics of tourism, and wildlife conservation in Africa. African wildlife foundation discussion papers series N.4

Emerton L (1998) Innovations for financing wildlife conservation in Kenya, presented at the 10 th Global Biodiversity forum, Bratislava

Emerton L (1999) Balancing the opportunity costs of wildlife conservation for communities around lake Mburo national park, Uganda

Emmanouilides CJ, Sgouromalli H (2013) Renewable energy sources in Crete: economic valuation results from a stated choice experiment. In 6 TH International Conference on Information And Communication Technologies In Agriculture, Food And Environment (Haicta 2013). Salampasis, M., Theodoridis, A. (Ed.) Book Series: Procedia Technology 8:406–415

Ericsson G, Kindberg J, Bostedt G (2007) Willingness to pay (WTP) for wolverine Gulo gulo conservation. Wildlife Biol 13:2–12

Everard M, Reed MS, Kenter JO (2016) The ripple effect: institutionalising pro-environmental values to shift societal norms and behaviours. Ecosyst Serv 21b:230–240

Farnsworth KD, Adenuga AH, de Groot RS (2015) The complexity of biodiversity: a biological perspective on economic valuation. Ecol Econ 120:350–354

Farnsworth KD, Lyashevska O, Fung T (2012) Functional complexity: the source of value in biodiversity. Ecol Complex 11:46–52

Farr M, Stoeckl N, Beg RA (2014) The non-consumptive (tourism) ‘value’ of marine species in the Northern section of the Great Barrier Reef. Mar Policy 43:89–103

Fischer A, Weldesemaet YT, Czajkowski M, Tadie D, Hanley N (2015) Trophy hunters’ willingness to pay for wildlife conservation and community benefits. Conserv Biol 29(4):1111–1121

Fix PJ, Manfredo MJ, Loomis JB (2005) Assessing validity of elk and deer license sales estimated by contingent valuation. Wildl Soc Bull 33(2):633–642

Forster J (ed) (1997) Valuing Nature? Routledge, New York

Frew KN, Peterson MN, Sills E, Moorman CE, Bondel H, Fueller JC, Howell DL (2018) Market and nonmarket valuation of North Carolina’s Tundra swans among hunters, wildlife watchers and the public. Wildl Soc Bull 42(3):478–487

Fried BM, Adams RM, Berrens RP, Bergland O (1995) Willingness to pay for a change in elk hunting quality. Wildl Soc Bull 23(4):680–686

Frontuto V, Dalmazzone S, Vallino E, Giaccaria S (2017) Earnmarking conservation: further inquiry on scope effects in stated preference method applied to nature-based tourism. Tour Manag 60:130–139

Gan F, Du H, Wei Q, Fan E (2011) Evaluation of the ecosystem values of aquatic wildlife reserves: a case of Chinese Sturgeon Natural Reserve in Yichang reaches of the Yangtzeriver. J Appl Ichthyol 27(2):376–382

Garrod GD, Willis KG (1994) Valuing biodiversity and nature conservation at a local-level. Biodivers Conserv 3(6):555–565

Garcia-Jimenez R, Morales-Reyes Z, Perez-Garcia JM, Margalida A (2021) Economic valuation of non-material contributions to people provided by avian scavengers: Harmonizing conservation and wildlife-based tourism. Ecol Econ 187:107088

Gascoigne W, Hill R, Haefele M, Loomis J, Hyberg S (2021) Economics of the Conservation Reserve Program and the wildlife it supports: a case study of upland birds in South Dakota. J Outdoor Recreat Tour 35:100385

Gavin MC, McCarte J, Berkes F, Mead ATP, Sterling EJ, Tang R, Turner NJ (2018) Effective biodiversity conservation requires dynamic, pluralistic, partnership-based approaches. Sustainability 10:1846. https://doi.org/10.3390/su10061846

Getzner M (2015) Importance of Free-Flowing Rivers for Recreation: Case Study of the River Mur in Styria, Austria. J Water Resour Plan Manag141(2): pages not reported. https://doi.org/10.1061/(ASCE)WR.1943-5452.0000442

Gómez-Baggethun E, Ruiz-Pérez M (2011) Economic valuation and the commodification of ecosystem services. Prog Phys Geogr 35:613–628

Gren IM, Haggmark-Svensonn T, Elofsson K, Engelmann M (2018) Economics of wildlife management. An Aoverview Eur J Wildl Res 64:22

Guimaraes MH, Nunes LC, Madureira L, Santos JL, Bosk T, Dentinho T (2015) Measuring birdwatchers preferences: a case for using online networks and mixed-mode surveys. Tour Manage 46:102–113

Han SY, Lee CK (2008) Estimating the value of preserving the Manchurian black bear using the contingent valuation method. Scand J for Res 23(5):458–465

Hanemann M (1994) Valuing the Environment through Contingent Valuation. JEP 8(4):19–43

Hanley N, Czajkowski M, Hanley-Nickolls R, Redpath S (2010) Economic values of species management options in human-wildlife conflicts Hen Harriers in Scotland. Ecol Econ 70(1):107–113

Hanley N, MacMillan D, Patterson I, Wright RE (2003) Economics and the design of nature conservation policy: a case study of wild goose conservation in Scotland using choice experiments. Anim Conserv 6:123–129

Hanley N, Hynes S, Patterson D, Jobstvogt N (2015) Economic Valuation of Marine and Coastal Ecosystems: is it currently fit for purpose? J Ocean Coast Econ 2(1):1–24

Harihar A, Verissimo D, MacMillan DC (2015) Beyond compensation: integrating local communities’ livelihood choices in large carnivore conservation. Global Environ Change 33:122–130

Hattam C, Böhnke-Henrichs A, Börger T, Burdon D, Hadjimichael M, Delaney A, Atkins JP, Garrard S, Austen MC (2015) Integrating methods for ecosystem service assessment and valuation: mixed methods or mixed messages? Ecol Econ 120:126–138

Hockley N (2014) Cost-benefit analysis: a decision-support tool or a venue for contesting ecosystem knowledge? Environ Plan C: Politics Space 32:283–300

Horne P, Petajisto L (2003) Preferences for alternative moose management regimes among Finnish landowners: a choice experiment approach. Land Econ 79(4):472–482

Howarth RB, Wilson MA (2006) A Theoretical Approach to Deliberative Valuation: Aggregation by Mutual Consent. Land Econ 82:1–16. https://doi.org/10.3368/le.82.1.1

Hawkins SA, Vangerschov Iversen S, Brady B, Mayhew M, Smith D, Lipscombe S, White C, Eagle A, Convery I (2020) Community perspectives on the reintroduction of Eurasian lynx (Lynx lynx) to the UK. Restor Ecol 28(6):1408–1418

Hutt CP, Hunt KM, Schlechte JW, Buckmeier DL (2013) Effects of catfish angler catch-related attitudes on fishing trip preferences. N Am J Fish Manag 33(5):965–976

Imamura K, Takano KT, Mori N, Nakashizuka T, Managi S (2016) Attitudes toward disaster-prevention risk in Japanese coastal areas: analysis of civil preference. Nat Hazards 82(1):209–226

IPBES (2019) Global assessment report on biodiversity and ecosystem services of the Intergovernmental Science - Policy Platform on Biodiversity and Ecosystem Services. E. S. Brondizio, J. Settele, S. Díaz, and H. T. Ngo (editors). IPBES Secretariat, Bonn, Germany

IPBES (2022) Summary for policymakers of the methodological assessment regarding the diverse conceptualization of multiple values of nature and its benefits, including biodiversity and ecosystem functions and services (assessment of the diverse values and valuation of nature) - Policy Platform on Biodiversity and Ecosystem Services. IPBES Secretariat, Bonn, Germany

Irvine KN, O’Brien L, Ravenscroft N, Cooper N, Everard M, Fazey I, Reed MS, Kenter JO (2016) Ecosystem services and the idea of shared values. Ecosyst Serv 21:184–193

Isacs L, Kenter J, Wetterstrand H, Katzeff C (2021) What does value pluralism mean in practice? An empirical demonstration from a deliberative valuation. People and Nature. In press

Jacobs S, Dendoncker N, Martín-López B, Barton DN, Gomez-Baggethun E, Boeraeve F, McGrath FL, Vierikko K, Geneletti D, Sevecke KJ, Pipart N, Primmer E, Mederly P, Schmidt S, Aragão A, Baral H, Bark RH, BricenoT BD et al (2016) A new valuation school: integrating diverse values of nature in resource and land use decisions. Ecosyst Serv 22:213–220

Jenkins H (2011) Opening address, 7th International Wildlife Ranching Symposium by Premier of the Northern Cape. Kimberly, South Africa: 7th International Wildlife Ranching Symposium

Jobstvogt N, Hanley N, Hynes S, Kenter JO, Witte U (2014a) Twenty thousand sterling under the sea: estimating the value of protecting deep-sea biodiversity. Ecol Econ 97:10–19. https://doi.org/10.1016/j.ecolecon.2013.10.019

Jobstvogt N, Watson V, Kenter JO (2014b) Looking below the surface: the cultural ecosystem service values of UK marine protected areas (MPAs ) . Ecosyst Serv 10:97–110. https://doi.org/10.1016/j.ecoser.2014.09.006

Johansson M, Sjostrom M, Karlsson J, Brannlund R (2012) Is human fear affecting public willingness to pay for the management and conservation of large carnivores. Soc Nat Resour 25(6):610–620

John KH, Youn YC, Shin JH (2003) Resolving conflicting ecological and economic interests in the Korean DMZ: a valuation based approach. Ecol Econ 46(1):173–179

Jones L, Norton L, Austin Z, Browne AL, Donovan D, Emmett BA, Grabowski ZA, Howard DC, Jones JPG, Kenter JO, Manley W, Morris C, Robinson A, Short C, Siriwardena GM, Stevens CL, Storkey J, Waters RD, Willis GF (2016) Stocks and flows of natural and human-derived capital in ecosystem services. Land Use Policy 52:151–162

Jordan T (2014) Deliberative methods for complex issues: a typology of functions that may need scaffolding. A Research and Applications Journal 13:50–71

Kaffashi S, Yacob MR, Clark M, S, Radam A, Mamat MF (2015) Exploring visitors’ willingness to pay to generate revenues for managing the national elephant conservation centre in Malaysia. For Policy Econ 56:9–19

Kahneman D (2011) Thinking Fast and Slow. Macmillan, New York

Kallis G, Gómez-Baggethun E, Zografos C (2013) To value or not to value? That is not the question. Ecol Econ 94:97–105

Kenter JO (2015) Valuing ecosystem services: what’s the value? Ecology 96:2848–2853

Kenter JO (2016a) Deliberative and Non-Monetary Valuation. in M. Potschin, R. Haines-Young, R. Fish, and R. K. Turner, editors. Routledge Handbook of Ecosystem Services. Routledge, Abingdon

Kenter JO (2016b) Editorial: shared, plural and cultural values. Ecosyst Serv 21:175–183

Kenter JO (2016c) Integrating deliberative monetary valuation, systems modelling and participatory mapping to assess shared values of ecosystem services. Ecosyst Serv 21:291–307

Kenter JO (2017) Deliberative Monetary Valuation. In C. L. Spash, editor. Routledge Handbook of Ecological Economics: Nature and Society . Routledge, Abingdon

Kenter JO, Reed M, Irvine K, O’Brien L, Brady E, Bryce R, Christie M, Church A, Cooper N, Davies A, Evely A, Everard M, Fazey I, Hockley N, Jobstvogt N., Molloy C, Orchard-Webb J, Ravenscroft N, Ryan M, Watson V (2014) UK National Ecosystem Assessment follow-on phase. Work Package Report 6: shared, plural and cultural values of ecosystems. UNEP-WCMC, Cambridge

Kenter JO, O’Brien L, Hockley N, Ravenscroft N, Fazey I, Irvine K, Reed MS, Christie M, Brady E, Bryce R, Church A, Cooper N, Davies A, Evely A, Everard M, Fish R, Fisher JA, Jobstvogt N, Molloy C et al (2015) What are shared and social values of ecosystems? Ecol Econ 111:86–99

Kenter JO, Bryce R, Christie M, Cooper N, Hockley N, Irvine K, Fazey I, O’Brien L, Orchard-Webb J, Ravenscroft N, Raymond C, Reed MS, Tett P, Watson V (2016a) Shared values and deliberative valuation: future directions. Ecosyst Serv 21:358–371

Kenter JO, Hyde T, Christie M, Fazey I (2011) The importance of deliberation in valuing ecosystem services in developing countries—evidence from the Solomon Islands. Glob Environ Change 21:505–521

Kenter JO, Jobstvogt N, Watson V, Irvine KN, Christie M, Bryce R (2016b) The impact of information, value-deliberation and group-based decision-making on values for ecosystem services: integrating deliberative monetary valuation and storytelling. Ecosyst Serv 21:270–290

Kenter JO, Reed MS, Fazey I (2016c) The deliberative value formation model. Ecosyst Serv 21b:194–207

Kenter JO, Raymond CM, Van Riper CJ, Azzopardi E, Brear MR, Calcagni F, Christie I, Christie M, Fordham A, Gould RK, Ives CD, Hejnowicz AP, Gunton R, Horcea-Milcu AI, Kendal D, Kronenberg J, Massenberg JR, O’Connor S, Ravenscroft N, Rawluk A, Raymond IJ, Rodríguez-Morales J et al (2019) Loving the mess: navigating diversity and conflict in social values for sustainability. Sustain Sci 14:1439–1461. https://doi.org/10.1007/s11625-019-00726-4

Kenter JO, O’Connor S (2021) The Life Framework of Values and living as nature; towards a full recognition of holistic and relational ontologies. Sustain Sci 17:2529–2542

Kenyon W, Nevin C (2001) The use of economic and participatory approaches to assess forest development: a case study in the Ettrick Valley. For Policy Econ 3(1–2):69–80

Knoche S, Lupi F (2013) Economic benefits of publicly accessible land for ruffed grouse hunters. J Wildl Manag 77(7):1294–1300

Kronenberg J (2015) Betting against human ingenuity: the perils of the economic valuation of nature’s services. Bioscience 65:1096–1099

Ku SJ, Yoo SH (2010) Willingness to pay for renewable energy investment in Korea: a choice experiment study. Renew Sust Energ Rev 14(8):2196–2201

Kubo T, Shoji Y (2014a) Spatial tradeoffs between residents’ preferences for brown bear conservation and the mitigation of human-bear conflicts. Biol Conserv 176:126–132

Kubo T, Shoji Y (2014b) Trade-off between human-wildlife conflict risk and recreation conditions. Eur J Wildl Res 60(3):501–510

Lawson T (2013) What is this “school” called neoclassical economics? Camb J Econ 37:947–983

Lienhoop N, Fischer A (2009) Can you be bothered? The role of participant motivation in the valuation of species conservation measures. J Environ Plan Manag 52(4):519–534

Lienhoop N, Bartkowski B, Hansjurgens B (2015) Informing biodiversity policy: the role of economic valuation, deliberative institutions and deliberative monetary valuation. Environ Sci Policy 54:522–532

Lienhoop N, MacMillan D (2007) Valuing wilderness in Iceland: Estimation of WTA and WTP using the market stall approach to contingent valuation. Land Use Policy 24:289–295

Lliso B, Marie, P, Pascual U, Engel S (2020) Increasing the credibility and salience of valuation through deliberation: lessons from the Global South. Glob Environ Change 62:102065. https://doi.org/10.1016/j.gloenvcha.2020.102065

Lo AY, Spash CL (2012). Deliberative monetary valuation: in search of a democratic and value plural approach to environmental policy. J Econ Surv 27:768–789. Wiley Online Library

Loomis JB (1993) Integrated Public Lands Management : Principles and Applications to National Forests, Parks, Wildlife Refuges and BLM Lands. Columbia University Press, New York, New York (USA)

Loomis JB (2000) Can environmental economic valuation techniques aid ecological economics and wildlife conservation? Wildl Soc Bull 28(1):52–60

Loomis JB, White DS (1996) Economic benefits of rare and endangered species: summary and meta-analysis. Ecol Econ 18:197–206

Lundhede TH, Jacobsen JB, Thorsen BJ (2015) A hedonic analysis of the complex hunting experience. J for Econ 21(2):51–66

Mace G (2014) Whose conservation? Science 345(6204):1558–1560

MacMillan DC, Philip L, Hanley N, Alvarez Farizo B (2002) Valuing the non-market benefits of wild goose conservation: a comparison of interview and group-based approaches. Ecol Econ 43:49–59

MacMillan DC, Hanley N, Lienhoop N (2006) Contingent valuation: Environmental polling or preference engine? Ecol Econ 60:299–407

Madhusudan MD (2003) Living amidst large wildlife: livestock and crop depredation by large mammals in the interior villages of Bhadra tiger reserve, South India. Environ Manage 31:0466–0475

Article   CAS   Google Scholar  

Manfredo MJ, Bruskotter JT, Tee TL, Fulton D, Schwartz SH, Arlinghaus R, Oishi S, Uskul AK, Redford K, Kitayama S, Sullivan L (2017) Why social values cannot be changed for the sake of conservation. Conserv Biol 31(4):772–780. https://doi.org/10.1111/cobi.12855

Manfredo MJ, Teel TL, Dietsch AM (2016) Implications of human value shift and persistence for biodiversity conservation. Conserv Biol 30:287–296

Margalida A, Donázar JA, Carrete M, Sánchez-Zapata JA (2010) Sanitary versus environmental policies: fitting together two pieces of the puzzle of European vulture conservation. J Appl Ecol 47:931–935. https://doi.org/10.1111/j.1365-2664.2010.01835.x

Margalida A, Colomer MA (2012) Modelling the effects of sanitary policies on European vulture conservation. Sci Rep 2:753. https://doi.org/10.1038/srep00753

Markandya A, Taylor T, Longo A, Murty MN, Murty S et al (2008) Counting the cost of vulture decline—an appraisal of the human health and other benefits of vultures in India. Ecol Econ 67:194–204. https://doi.org/10.1016/j.ecolecon.2008.04.020

Martínez-Alier J, Munda G, O’Neill J (1998) Weak comparability of values as a foundation for ecological economics Ecol . Econ 26(3):277–286

Martinez-Espineira R (2007) ‘Adopt a hypothetical pup’: a count data approach to the valuation of wildlife. Environ Resour Econ 37(2):335–360

Martinez-Espineira R (2006) A Box-Cox Double-Hurdle model of wildlife valuation: the citizen’s perspective. Ecol Econ 58(1):192–208

Martin-Lopez B, Montes C, Benayas J (2007) Economic valuation of biodiversity conservation: the meaning of numbers. Conserv Biol 22(3):624–635

Martino S, Tett P, Kenter JO (2019) The interplay between economics, legislative power and social influence examined through a social-ecological framework for marine ecosystems services. Sci Total Environ 651(1):1388–1404

Martino S, Kenter J, Albers N, Whittingham MJ, Young DM, Pearce-Higgins JW, Martin-Ortega J, Glenk K, Reed MS (2022) Trade-offs between the natural environment and recreational infrastructure: a case study about peatlands under different management scenarios. Land Use Policy 23:106401

Mascia MB, Brosius JP, Dobson TA, Forbes BC, Horowitz L, McKean MA, Turner NJ (2003) Conservation and social sciences. Conserv Biol 17:649–650

Matulis BS (2014) The economic valuation of nature: a question of justice? Ecol Econ 104:155–157. https://doi.org/10.1016/j.ecolecon.2014.04.010

McCauley DJ (2006) Selling out on nature. Nature 443:27–28

McNeely J (1993) Economic incentives for conserving biodiversity: lessons from Africa. Ambio 22(2/3):144–150

MEA-Millennium Ecosystem Assessment (2005) Ecosystems and human well-being: biodiversity synthesis. World Resources Institute, Washington, D.C.

Meffe GK, Caroll CR (1994) Principles of Conservation Biology. Sinauer

Mensah JT, Elofsson K (2017) An empirical analysis of hunting lease pricing and value on game in Sweden. Land Econ 93(2):292–308

Mitchell RC, Carson R (1989) Using Surveys to Value Public Goods: The Contingent Valuation Method. Resources for the Future, Washington, DC

Mmopelwa G, Kgathi DL, Molefhe L (2007) Tourists’ perceptions and their willingness to pay for park fees: a case study of self-drive tourists and clients for mobile tour operators in Moremi Game Reserve. Botswana Tour Manag 28(4):1044–1056

Molina JR, Zamora R, Rodríguez y Silva F (2019) The role of flagship species in the economic valuation of wildfire impacts: an application to two Mediterranean protected areas. Sci Total Environ 675:520–530

Morales-Reyes Z, Pérez-García JM, Moleón M, Botella F, Carrete M et al (2015) Supplanting ecosystem services provided by scavengers raises greenhouse gas emissions. Sci Rep 5:7811. https://doi.og/ https://doi.org/10.1038/srep07811

Moro M, Fischer A, Milner-Gulland EJ, Lowassa A, Naiman LC, Hanley N (2015) A stated preference investigation of household demand for illegally hunted bushmeat in the Serengeti. Tanzania Anim Conserv 18(4):377–386

Morse-Jones S, Bateman IJ, Kontoleon A, Ferrini S, Burgess ND, Turner RK (2012) Stated preferences for tropical wildlife conservation amongst distant beneficiaries: Charisma, endemism, scope and substitution effects. Ecol Econ 78:9–18

Muchapondwa E, Carlsson F, Kohlin G (2008) Wildlife management in Zimbabwe: evidence from a contingent valuation study. S Afr J Econ 76(4):685–704

Navrud S, Mungatana ED (1994) Environmental valuation in developing countries: the recreational value of wildlife viewing. Ecol Econ 11:135–151

Nguyen VV, Chun-Hung L, Phan TTT (2022) Integrating multiple aspects of human-elephant conflict management in Dong Nai Biosphere Reserve. Vietnam Glob Ecol Conserv 39:e02285

Nea UK (2011) The UK National Ecosystem Assessment. CUP, Cambridge

Nielsen ASE, Lundhede TH, Jacobsen JB (2016) Local consequences of national policies - A spatial analysis of preferences for forest access reduction. For Policy Econ 73:68–77

Nielsen MR, Jacobsen JB, Thorsen BJ (2014) Factors Determining the Choice of Hunting and Trading Bushmeat in the Kilombero Valley. Tanzania Conserv Biol 28(2):382–391

Ninan KN, Sathyapalan J (2005) The economics of biodiversity conservation: a study of a coffee growing region in the Western Ghats of India. Ecol Econ 55(1):61–72

Nordhaus W (2017) Revisiting the social cost of carbon. PNAS 114(7):1518–1523

Norgaard RB (2010) Ecosystem services: From eye-opening metaphor to complexity blinder. Ecol Econ 69:1219–1227

Nunes Paulo ALD, van der Bergh Jeroen CJM (2001) Economic valuation of biodiversity: sense or nonsense? Ecol Econ 39:203–222

Nuno A, Blumenthal JM, Austin TJ, Bothwell J, Ebanks-Petrie G, Godley BJ, Broderick AC (2018) Understanding implications of consumers behaviour for wildlife farming and sustainable wildlife trade. Conserv Biol 32(2):390–400

O’Brien EA (2003) Human values and their importance to the development of forestry policy in Britain: a literature review. Forestry 76:3–17

O’Connor S, Kenter JO (2019) Making intrinsic values work; integrating intrinsic values of the more-than-human world through the Life Framework of Values. Sust Sci 14:1247–1265. https://doi.org/10.1007/s11625-019-00715-7

OECD (2013) Scaling Up Finance Mechanisms for Biodiversity

Ojea E, Loureiro ML (2007) Altruistic, egoistic and biospheric values in willingness to pay (WTP) for wildlife. Ecol Econ 63(4):807–814

O’Neill J, Holland A, Light A (2008) Environmental values. Routledge, Abingdon

Book   Google Scholar  

Orchard-Webb J, Kenter JO, Bryce R, Church A (2016) Deliberative Democratic Monetary Valuation to implement the Ecosystem Approach. Ecosyst Serv 21:308–318

Ostrom E (1990) Governing the common. Cambridge University Press

Owen R, Duinker P, Beckley T (2009) Capturing old-growth values for use in forest decision-making. Environ Manage 43(2):237–248

Parks S, Gowdy J (2013) What have economists learned about valuing nature? A Review essay. Ecosyst Serv 3:e1–e10

Pascual U, Balvanera P, Díaz S, Patak G, Roth E, Stenseke M, Watson RT, Başak Dessane E, Islar M, Kelemen E, Maris V, Quaas M, Subramanian SM, Wittmer H, Adlan A, Ahn S, Al-Hafedh YS, Amankwah E, Asah ST, Berry P, Bilgin A, Breslow SJ, Bullock C, Cáceres D, Daly-Hassen H, Figueroa E, Golden CD, Gómez-Baggethun E, González-Jiménez D, Houdet J, Keune H, Kumar R, Ma K, May PH, Mead A, O’Farrell P, Pandit R, Pengue W, Pichis-Madruga R, Popa F, Preston S, Pacheco-Balanza D, Saarikoski H, Strassburg BB, van den Belt M, Verma M, Wickson F, Yagi N (2017) Valuing nature’s contributions to people: the IPBES approach. Curr Opin Environ Sustain 26–27:7–16. https://doi.org/10.1016/j.cosust.2016.12.006

Pate J, Loomis J (1997) The effect of distance on willingness to pay values: a case study of wetlands and salmon in California. Ecol Econ 20(3):199–207

Pearce D, Turner RK (1989) Economics of natural resources and the environment. Pearson Education Limited

Pearce D, Moran D (1994) The economic value of biodiversity. IUCN, Earthscan Publications Limited, London

Pearce D, Atkinson G, Mourato S (2006) Cost –benefit analysis and the environment. recent developments. OECD, pp. 314

Perman R, Ma Y, McGilvray Y, Common M (2003) Natural Resource and Environmental Economics, 3rd edn. Pearson Addison Wesley, New York

Petrolia DR, Interis MG, Hwang J (2014) America’s Wetland? A National Survey of Willingness to Pay for Restoration of Louisiana’s Coastal Wetlands. Mar Resour Econ 29(1):17–37

Philips L, MacMillan D (2005) Exploring values, context and perceptions in contingent valuation studies: the CV Market Stall Technique and Willingness to Pay for Wildlife Conservation. J Environ Plan Manag 48(2):257–274

Pimentel D, Bailey O, Kim P, Mullaney E, Calabrese J, Walman L, Nelson F, Yao X (1999) Will limits of the Earth’s resources control human numbers? Environ Dev Sustain 1(1):19–39

Pires NM, Garla RC, Carvalho AR (2016) The economic role of sharks in a major ecotourism archipelago in the western South Atlantic. Mar Policy 72:31–39

Ranger S, Kenter JO, Bryce R, Cumming G, Dapling T, Lawes E, Richardson PB (2016) Forming shared values in conservation management: an interpretive-deliberative-democratic approach to including community voices. Ecosyst Serv 21:344–357

Rathnayake RMW (2016) Pricing the enjoyment of ‘elephant watching’ at the Minneriya National Park in Sri Lanka: An analysis using CVM. Tour Manag Perspect 18:26–33

Ravenscroft N (2010) The mythologies of environmental economics. Journal of Policy Research in Tourism, Leisure and Events 2:129–143

Raymond CM, Kenter JO, Plieninger T, Turner NJ, Alexander KA (2014) Comparing instrumental and deliberative paradigms underpinning the assessment of social values for cultural ecosystem services. Ecol Econ 107:145–156

Raymond CM, Kenter JO (2016) Transcendental values and the valuation and management of ecosystem services. Ecosyst Serv 21:241–257

Redpath SM, Young J, Evely A, Adams WM, Sutherland WJ, Whitehouse A, Amar A, Lambert RA, Linnell J, Watt A, Gutiérrez RJ (2013) Understanding and managing conservation conflicts. Trends Ecol Evol 28:100–109

Richardson L, Loomis J (2009) The total economic value of threatened, endangered and rare species: an updated meta-analysis. Ecol Econ 68:1535–1548

Richardson L, Rosen T, Gunther K, Schwartz C (2014) The economics of roadside bear viewing. J Environ Manage 140:102–110

Richardson L, Lewis L (2022) Getting to know you: individual animals, wildlife webcams, and willingness to pay for brown bear preservation. Am J Agric Econ 104(2):673–692

Rodriguez LC, Henson D, Herrero M, Nkedianye D, Reid R (2012) Private farmers’ compensation and viability of protected areas: the case of Nairobi National Park and Kitengela dispersal corridor. Int J Sustain Dev World Ecol 19(1):34–43

Roughgarden J (1995) Can economics save biodiversity? In: Swanson T (ed) The economics and ecology of biodiversity decline: the forces driving global change. Cambridge University Press, New York, pp 149–153

Saayman M, Saayman A (2017) Is the rhino worth saving? A sustainable tourism perspective. J Sustain Tour 25(2):251–264

Saayman M (2014) The non-consumptive value of selected marine species at Table Mountain national park: an exploratory study. SAJEMS 17(2):184–193

Saayman M, Saayman A (2014) Who is willing to pay to see the Big 7? Tour Econ 20(6):1181–1198

Sagoff M (1998) Aggregation and deliberation in valuing environmental public goods: a look beyond contingent pricing. Ecol Econ 24:213–230

Sandbrook C, Adams W, Büscher B, Vira B (2013) Social Research and Biodiversity Conservation. Conserv Biol 27(6):1487–1490. http://doi.wiley.com/ https://doi.org/10.1111/cobi.12141

Sandhu HS, Wratten SD, Cullen R, Case B (2008) The future of farming: the value of ecosystem services in conventional and organic arable land. An Experimental Approach Ecol Econ 64:835–848. https://doi.org/10.1016/j.ecolecon.2007.05.007

Sandhu H, Wratten S, Costanza R, Pretty J, Porter JR, Reganold J (2015) Significance and value of non-traded ecosystem services on farmland. Peer J.  https://doi.org/10.7717/peerj.762

Saunders CD, Brook AT, Eugene Myers O (2006) Using psychology to save biodiversity and human well-being. Conserv Biol 20:702–705

Scholte SSK, van Teeffelen AJA, Verburg PH (2015) Integrating socio-cultural perspectives into ecosystem service valuation: a review of concepts and methods. Ecol Econ 114:67–78

Schröter M, Zanden EH, Oudenhoven APE, Remme RP, Serna Chavez HM, Groot RS, Opdam P (2014) Ecosystem services as a contested concept: a synthesis of critique and counter-arguments. Conserv Lett 7:514–523

Sementelli A, Smith HT, Meshaka W, Engeman R (2008) Just Green Iguanas? The Associated Costs and Policy Implications of Exotic Invasive Wildlife in South Florida. USDA National Wildlife Research Center - Staff Publications. 1036. https://digitalcommons.unl.edu/icwdm_usdanwrc/1036

Serenari C, Shaw J, Myers R, Cobb DT (2019) Explaining deer hunter preferences for regulatory changes using choice experiments. J Wildl Manag 83(2):446–456

Shafer EL, Carline R, Guldin RW, Cordell HK (1993) Economic amenity values of wildlife - 6 case-studies in Pennsylvania. Environ Manage 17(5):669–682

Schaafsma M, Bartkowski B, Lienhoop N (2018) Guidance for deliberative monetary valuation studies. IRERE 12:267–323. https://doi.org/10.1561/101.00000103

Shapansky B, Adamowicz W, Boxall P (2003) Measuring forest resource values: an assessment of choice experiments and preferences construction methods as public involvement tools. Rural Economy, Project Report 02–03. University of Alberta, Edmonton, Canada

Shogren JF, Tschirhart J, Anderson T, Whritenour Ando A, Beissinger SR, Brookshire D, Brown GM Jr, Coursey D, Innes R, Meyer SM, Polasky S (1999) Why economics matters for endangered species protection. Conserv Biol 13(6):1257–1261

Shwiff SA, Anderson A, Cullen R, White PCL, Shwiff SS (2013) Assignment of measurable costs and benefit to wildlife conservation projects. Wild Res 40:134–141

Siachoono SM (1995) Contingent valuation as an additional tool for evaluating wildlife utilization management in Zambia – Mumbwa game management area. Ambio 24(24):246–249

Sinclair M, Ghermandi A, Sheela AM (2018) A crowdsources valuation of recreational ecosystem services using social media data: an application to a tropical wetland in India. Sci Tot Environ 642:356–365

Spash CL, Aslaksen I (2015) Re-establishing an ecological discourse in the policy debate over how to value ecosystems and biodiversity. J Environ Manage 159:245–253

Spash CL, Vatn A (2006) Transferring environmental value estimates: issues and alternatives. Ecol Econ 60:379–388

Spash CL (2006) Non-economic motivation for contingent values: rights and attitudinal beliefs in the willingness to pay for environmental improvements. Land Econ 82:602–622. University of Wisconsin Press

Spash CL (2008a) Deliberative monetary valuation and the evidence for a new value theory. Land Econ 84:469–488

Spash CL (2008b) How much is that ecosystem in the window? The one with the bio-diverse trail. Environ Values 17(2):259–284. https://doi.org/10.3197/096327108x303882

Spash CL (2007) Deliberative monetary valuation (DMV): issues in combining economic and political processes to value environmental change. Ecol Econ 63:690–699

Spash CL (2002) Informing and forming preferences in environmental valuation: coral reef biodiversity. J Econ Psychol 23:665–687

Spash CL (2015) Bulldozing biodiversity: the economics of offsets and trading-in Nature. Biol Conserv 192:541–551. https://doi.org/10.1016/j.biocon.2015.07.037

Spenceley A, Rylance A, Laiser SL (2017) Protected area entrance fee in Tanzania: the search for competitiveness and value for money. Koedoe 59(1):article number a1442

Stanley DL (2005) Local perception of public goods: recent assessments of willingness-to-pay for endangered species. Contemp Econ Policy 23(2):165–179

Stefanski SE, Villasante S (2015) Whales vs. gulls: assessing trade-offs in wildlife and waste management in Patagonia. Argentina Ecosyst Serv 16:294–305

Sterner RT (2009) The economics of threatened species conservation: a review and analysis. National Wildlife Research Center - Staff Publications, Paper, p 978

Sutton WR, Larson DM, Jarvis LS (2008) Assessing the costs of living with wildlife in developing countries using willingness to pay. Environ Dev Econ 13:475–459

Szabó Z (2011) Reducing protest responses by deliberative monetary valuation: Improving the validity of biodiversity valuation. Ecol Econ 72:37–44

Tait P, Saunders C, Nugent G, Rutherford P (2017) Valuing conservation benefits of disease control in wildlife. A choice experiment approach to bovine tuberculosis management in New Zealand native forests. J Environ Manage 189:142–149

Tallis H, Kareiva P, Marvier M, Chang A (2008) An ecosystem services framework to support both practical conservation and economic development. PNAS 105:9457–9464

Teal GA, Loomis JB (2000) Effects of gender and parental status on the economic valuation of increasing wetlands, reducing wildlife contamination and increasing salmon populations. Soc Nat Resour 13(1):1–14

TEEB (2010) The economics of ecosystem services and biodiversity. Mainstreaming the economics of nature: a synthesis of the approach, conclusion and recommendations

Tengö M, Brondizio ES, Elmqvist T, Malmer P, Spierenburg M (2014) Connecting diverse knowledge systems for enhanced ecosystem governance: The multiple evidence base approach. Ambio 43:579–591. Springer Netherlands

Tisdell C, Wilson C (2004) The public’s knowledge of and support for conservation of Australia’s tree-kangaroos and other animals. Biodivers Conserv 13(12):2339–2359

Tisdell C, Swarna NH, Wilson C (2007) Endangerment and likeability of wildlife species: how important are they for payments proposed for conservation? Ecol Econ 60(3):627–633

Tisdell C, Wilson C, Swarna Nantha H (2005a) Policies for saving a rare Australian glider: economics and ecology. Biol Conserv 123:237–248

Tisdell C, Wilson C, Swarna NH (2005b) Association of public support for survival of wildlife species with their likeability. Anthrozoos 18(2):160–174

Travers H, Archer LJ, Mwedde G, Roe D, Baker J, Plumptre AJ, Rwetsiba A, Milner-Gulland EJ (2019) Understanding complex drivers of wildlife crime to design effective conservation interventions. Conserv Biol 33(6):1296–1306

Turner RK (2016) The “Balance Sheet” approach within adaptive management for ecosystem services. In: Potschin M, Haines-Young R, Fish R, Turner RK (eds) Routledge Handbook of Ecosystem Services. Routledge, London, pp 289–303

Chapter   Google Scholar  

Turner K, Paavola J, Cooper P, Jessamy V, Rosendo S, Georgiou S (2003) Valuing nature: lessons learned and future research directions. Ecol Econ 46:493–510

Turner RK (2007) Limits to CBA in UK and European environmental policy: retrospect and future prospects. CSERGE Working Paper EDM 06–17. ISSN 0967–8875

Turnhout E, Waterton C, Neves K, Buizer M (2013) Rethinking biodiversity: from goods and services to “living with.” Conserv Lett 6:154–161

Van der Bergh J, Botzen W (2014) A lower bound to the social cost of CO2 emissions. Nat Clim Change 4:253–258

van Beukering P, Brander L, Immerzeel D (2014) Value after the volcano: economic valuation of Montserrat’s Centre Hills. In K. N. Ninan, editor. Valuing Ecosystem Service: Methodological Issues and Case Studies. Edward Elgar, Cheltenham

van Eeden LM, Bogezi C, Leng DY, Marzluff JM, Wirsing AJ, Rabotyagov S (2021) Public willingness to pay for gray wolf conservation that could support a rancher-led wolf-livestock coexistence program. Biol Conserv 260:109226.  https://doi.org/10.1016/j.biocon.2021.109226

Vargas A, Diaz D (2017) Going along with the crowd? The importance of group effects for environmental deliberative monetary valuation. Cuadernos De Economía 36(70):75–94

Vargas A, Lo AY, Rohde N, Howes M (2017) Social influences on expressed willingness to pay: results of a deliberative monetary valuation study in Colombia. J Environ Plan Manag 60(9):1511–1528

Völker M, Lienhoop N (2016) Exploring group dynamics in deliberative choice experiments. Ecol Econ 123:57–67

Wakamatsu M, Shin KJ, Wilson C, Managi S (2018) Exploring a gap between Australia and Japan in the economic valuation of whale conservation. Ecol Econ 146:397–407

Wang KQ, Liu HM, Hu WY, Cox L (2016) Using online self-assessment tool to improve conjoint analysis Application in choices of wildlife excursions. Internet Res 26(3):644–660

Wang ZZ, Gong YZ, Mao XQ (2018) Exploring the value of overseas biodiversity to Chinese netizens based on willingness to pay for the African elephants’ protection. Sci Tot Environ 637:600–608

Ward H (1999) Citizens’ juries and valuing the environment: a proposal. Env Polit 8(2):75–96

Wätzold F, Lienhoop N, Drechsler M, Settele J (2008) Estimating optimal conservation in the context of agri-environmental schemes. Ecol Econ 68:295–305

White C, Convery I, Eagle A, O’Donoghue P, Piper S, Rowcroft P, Smith D, van Maanen E (2015) Cost-benefit analysis for the reintroduction of lynx to the UK: main report. AECOM

Whitehead JC, Aiken R (2007) Temporal reliability of willingness to pay from the National Survey of Fishing. Hunting and Wildlife-Associated Recreation Appl Econ 39(4–6):777–786

Whitten SM, Bennett JW (2002) A travel cost study of duck hunting in the Upper South East of South Australia. Aust Geog 33(2):207–221

Yoon CG (2009) Wise use of paddy rice fields to partially compensate for the loss of natural wetlands. Paddy Water Environ 7(4):357–366

Zambrano-Monserrate MA (2020) The economic value of the Andean Condor: the national symbol of South America. J Nat Conserv 54:125796. https://doi.org/10.1016/j.jnc.2020.125796

Zawacki WT, Marsinko A, Bowker JM (2000) A travel cost analysis of non-consumptive wildlife-associated recreation in the United States. For Sci 46(4):496–506

Zimmerhackel JS, Rogers SS, Meekan MG, Ali K, Pannell DJ, Kragt ME (2018) How shark conservation in the Maldives affects demand for dive tourism. Tour Manag 69:263–271

Zimmermann A, Albers N, Kenter JO (2021) Deliberating our frames: how members of multi-stakeholder initiatives use shared frames to tackle within-frame conflicts over sustainability issues. J Bus Ethics. https://doi.org/10.1007/s10551-021-04789-1

Zhou X, MacMillan DC, Zhang W, Wang Q, Jin Y, Verissimo D (2021) Understanding the public debate about trophy hunting in China as a rural development mechanism. Anim Conserv 24(3):346–354

Zsabo Z (2011) Reducing protest responses by deliberative monetary valuation: Improving the validity of biodiversity valuation. Ecol Econ 72:37–44

Download references

Author information

Simone Martino and Jasper O. Kenter have contributed equally.

Authors and Affiliations

Department of Environment and Geography, University of York, York, YO10 5NG, UK

Simone Martino & Jasper O. Kenter

The James Hutton Institute, AB15 8QH, Craigiebuckler, Aberdeen, Scotland

Simone Martino

Ecologos Research, Borth, SY245JQ, Wales, UK

Jasper O. Kenter

University of Aberystwyth - Business School, Penglais Campus, Aberystwyth University, Hugh Owen Building, Aberystwyth, SY23 3DY, Ceredigion, UK

You can also search for this author in PubMed   Google Scholar

Corresponding author

Correspondence to Simone Martino .

Ethics declarations

Competing interests.

The authors declare no competing interests.

Additional information

Publisher's note.

Springer Nature remains neutral with regard to jurisdictional claims in published maps and institutional affiliations.

Supplementary Information

Below is the link to the electronic supplementary material.

Supplementary file1 (DOCX 87 KB)

Rights and permissions.

Springer Nature or its licensor (e.g. a society or other partner) holds exclusive rights to this article under a publishing agreement with the author(s) or other rightsholder(s); author self-archiving of the accepted manuscript version of this article is solely governed by the terms of such publishing agreement and applicable law.

Reprints and permissions

About this article

Martino, S., Kenter, J.O. Economic valuation of wildlife conservation. Eur J Wildl Res 69 , 32 (2023). https://doi.org/10.1007/s10344-023-01658-2

Download citation

Received : 14 May 2022

Revised : 30 January 2023

Accepted : 10 February 2023

Published : 10 March 2023

DOI : https://doi.org/10.1007/s10344-023-01658-2

Share this article

Anyone you share the following link with will be able to read this content:

Sorry, a shareable link is not currently available for this article.

Provided by the Springer Nature SharedIt content-sharing initiative

• Ecological economics
• Environmental valuation
• Shared values
• Social values
• Value plurality
• Conservation values
• Find a journal
• Publish with us
• Track your research