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Biodiversity and Protected Areas exist neither in isolation nor independent of human activities. For local communities, this may mean conservation represents a hindrance rather than an opportunity for sustainable development and thus lead to increasing avoidance of the regulatory framework in effect. This paper defends changes to conservation practices in order to create a broader consensus around objectives and practices. One means of doing this is to ensure people adopt profit-based conservation practices. We discuss the advantages and disadvantages of economic valuation as a framework incentive measure to enforce local co-operation in conservation decisions and management. By using a methodological and conceptual approach, we seek to assess the reasons economic valuation, albeit an abstract, very theoretical and technical demanding indicator, may still be a useful conservation tool serving as an incentive and support to decision-making, as a tool in education and a vehicle of information.

Introduction

Traditional biological conservation management practices have been based on protected areas (PAs)1 (IUCN 2003) and the Safe Minimum Standard Principle (SMS)2 in line with the ecological perspective of sustainability as defined by the Daily Rule3 (Goodstein 1999). This has proved incapable of achieving the required conservation levels for PAs with resident populations, located near urban centres and with good means of access and especially within a low or middle income country context (OECD 1996). One reason for such relative failure is that biodiversity in general and PAs in particular exist neither in isolation nor are independent of human activity. Over 50% of the 20,000

1 A Protected Area (PA) is an ¡°area of land and/or sea especially dedicated to the protection and maintenance of biological diversity, and of natural and associated cultural resources, and managed through legal or other effective means¡± (IUCN 1994). In this paper biological diversity means ¡°¡¦the dynamic network of biological, chemical, and physical interactions that sustain a community and allow it to respond to changes in environmental conditions¡± (in OECD 1999, p. 41). Consequently PA means biological diversity conservation, or ecosystem conservation as a whole, with their geographical, biological and different types of resilience to different types of human activities resulting pressures. By doing so we will be following the Subsidiary Body on Scientific, Technical and Technological Advice of the Conference of the Parties to the Convention on Biological Diversity that indicates the importance of looking at biodiversity conservation under an ecosystem approach rather than focusing on the individual components within the ecosystems.
2 The SMS principle states that a sufficient area of ecosystem must be conserved to ensure the continued provision of ecosystem services unless the social costs of conservation are unacceptably high.
3 The Daily Rule states ¡°never reduce the stock of natural capital below a level that generates a sustained yield unless good substitutes are currently available to the services generated¡± (in Goodstein 1999, p. 112, citing Daly H. 1996. Beyond Growth: The Economics of Sustainable Development: 81-82. Beacon Press: Boston).

official PAs established over the last 200 years are on land historically occupied by indigenous peoples (IUCN 2003). Furthermore, some stress that official conservation initiatives have tended to neglect some of the subsequently protected environments as the result of their long-term interaction with humans1. In Europe, and in Portugal in particular, the ecosystems richest in biodiversity, such as alpine meadows or estuary zones, are those directly modified by humans and farmed animals.

Despite the richness and historical record of such human environmental interaction, modern conservation practices have implemented an ideology primarily based on nature and some kind of ¡°conservation ethic¡± potentially devaluing those values and systems that sustained human ecological practices in their respective contexts. While some governmental conservation agencies broadly recognise and accept the notion of ¡°conservation-with-use¡±2 by explicitly integrating it into national conservation legislation, most state-declared PA local communities have not been active participants in designating or managing their surroundings. This state-imposed development and conservation, leaving local community on the peripheries of power, has created local antagonism and suspicion (IUCN 2003). PA inhabitants, be they local farmers, herders or fisherman, all complain about being marginalised. For instance, in Portugal, where conservation policy is highly government-centralised, there is a strong feeling in some PAs that conservation serves only the interests of politicians, the urban elite and stakeholder interests rather than genuine local aspirations. When the state sets aside PAs, residents claim it does so at the cost of their own land, resources, and livelihoods, mostly without proper compensation for loss of property or forgone revenues (Wells and Brandon 1992). The act of conservation is thus interpreted as a misfortune rather than an opportunity for sustainable development. This thereby leads to increasing evasion of conservation regulations while the government delays in answering the worsening conservation problems resulting from those individual actions (McNeely 1988, Constanza et al 1997).

In this paper, we firmly support PAs as an insurmountable command-and-control instrument leading to the maintenance of biodiversity and ecological processes and life-support systems. Furthermore, current economic investigation is proving that enforced resource conservation measures are efficient in the sense that they do not cause Pareto inefficiency and that Pareto optimal allocations cannot only be reached through competitive markets (Gerlagh and Keyzer 2002). However, we would additionally state conservation practices have to change sharply. They must stretch beyond the SMS principle through both engaging locals and other users with the conservation process and creating a broad consensus as to the existence and objectives of conservation initiatives. One means of achieving this is to make people aware of the real benefits of the PA and ensure they understand they can benefit, even monetarily, from land being set aside by government for conservation purposes.

There are alternative forms of encouraging society to implement the Daily Rule beyond the SMS. Since economic decisions are market-based, it may be necessary to apply economic incentives to

1 This is surely the case of Europe, and Portugal in particular. In these regions, landscapes encompass large areas of semi-natural vegetation interspersed with grazing areas, hedgerows, farmland, and small villages and towns. Or, as in the case of wetlands, the coastline is frequently host to important urban communities depending on fishing activities and/or other marine activities and/or tourism.
2 IUCN 2003, p. 8.

conservation policy. The key idea is to ensure PA inhabitants, the owners of the natural capital1, gain a strong economic interest in protecting ecosystem sustainability either for their own profit or to maintain resale value (Goodstein 1999). This is profit-based conservation and we defend its application in conjunction with the SMS principle to improve conservation policy rather than as any simplistic alternative2. Within this framework, Co-Managed Protected Areas (CMPA) are emerging as a universally accepted means of community enforcement (IUCN 2003). CMPA are ¡°official state-established PAs managed with the effective engagement of other social actors, including indigenous and local communities¡± (IUCN 2003, p. 22).

The CMPA model has three objectives: the conservation of local natural and cultural heritage, the participation of civil society in the management process and the equitable distribution of benefits and costs. Community empowerment has to be reinforced with the implementation of incentive measures to enforce actor compliance, reinforce capacity building and provide intelligible information about the value of nature conservation. This is by no means an easy task for policy-makers and is packed with social, scientific, and practical difficulties and ambiguities (Heal 2000, Metrick and Weitzman 2000). Difficulties arise in establishing practical, common working definitions for indicators on biodiversity management such as diversity, ecosystem integrity, ecosystem health, sustainability resilience and species. There is sometimes a transnational dimension as appropriate solutions may cross geo-political frontiers requiring multi-lateral co-operation. A whole range of actors is affected and affects the existence or loss of nature (conservationists, local populations, entrepreneurs, policy-makers, etc.).

Furthermore, biodiversity policy has its own respective complexities, differing to classical pollution problems, including heterogeneity, irreversibility, accumulation of impacts, information gaps, mix of values and pressures (OECD 1999).

There are several incentives possible to enforce users3 and local community conservation compliance4. In this paper, we discuss the advantages and disadvantages of economic valuation of ecosystems as an incentive measure to enforce local community co-operation in conservation decisions and management. This issue is not new to economic and ecological literature. However, the literature mostly serves to demonstrate that economic valuation is anything but exhausted as an issue for discussion. Misleading pro and contra arguments and definitions are often used, both by economists and ecologists, making any understanding of monetary evaluation of biodiversity and its respective methodology more complicated than is justifiable. By recognising such difficulties, this paper tries to assess the main points of discussion around the economic valuation of biodiversity, including the advantages and disadvantages of applying it as an incentive tool to enforce conservation attitudes. Given the scope is rather extensive, containing highly controversial points of view, analysis is only provided from the economist¡¯s point of view and does not seek to be exhaustive. The main

1 Natural capital is generally defined as the stock of environmentally provided assets (e.g. soil, atmosphere, forests, water, wetlands, minerals) that provide flows of goods and services that are appropriate by economic sector and society at free cost (Serageldin 1996)
2 Various economists and ecologists have been warning that too little nature will be conserved by market force alone (e.g. Jakobsson and Dragun 1996, Constanza et al. 1997).
3 See Mendes 2003 for a theoretical review and travel cost empirical approach to estimate the value of an entrance fee for a National Park.
4 See OECD 1999 and McNeely 1988 for a comprehensive description about incentive measures for biodiversity.

objective is to provide an economics based opinion as to the most controversial aspects of ecosystem conservation.

The paper is organised as follows. We proceed by clarifying what ¡°value¡± means to economics. In section 3, we describe the most important steps required for ascertaining the most accurate possible monetary value. In section 4, we describe why monetary valuation is an important and reliable tool in policy and conservation management despite these difficulties and controversies. Finally, conclusions are drawn.

1. The Value of Ecosystems: What Does this Mean to Economics?1

In common usage, value means importance or desirability. To an economist, the value of an ecosystem is related to the contribution it makes to human wellbeing2. We are dealing with a very clear anthropocentric, utilitarian viewpoint according to which ecosystems are valuable insofar as they serve humans or to the extent they confer satisfaction on humans (Goulder and Kennedy 1997). We would like to underline that utilitarianism is not necessarily synonymous with exploitation or depletion of nature. On the contrary, it can be consistent with nature conservation where protection is perceived as a source of satisfaction or wellbeing.

The utilitarian approach allows value to arise in a number of ways depending on how individuals use ecosystems. The prior value of the ecosystem, also called primary value, consists of the system characteristics upon which all ecological functions depend (resilience capacity, individual resource stability, biodiversity retention) (Turner 1999). Their value arises in the sense that they produce other functions with value . secondary functions. These secondary functions and associated values depend on the maintenance, health, existence and the operationality of the ecosystem as a whole. The primary value is however related to the fact that the ecosystem holds secondary functions and values and, as such, in principle, has economic value.

Hence, economists have generally settled for taxonomy of total ecosystem value interpreted as a Total Economic Value (TEV), that distinguishes between Direct Use Values and Passive3 (Non-use) Values. TEV and its components have been the subject of huge debate among environmental economists, ecologists, psychologists, and others, about the viability, the usefulness, or the ethics of monetising it, especially passive uses4. Nevertheless there is actually a growing trend towards using the TEV measure on the grounds that theoretically there is no need to adopt a dichotomy that involves the adoption of arbitrary assumptions5. Advances in ecological economic models and theory also seem to stress the value of the overall system6 as opposed to individual system components. This points to the value of the system itself when exhibiting resilience capacity defined as the ability of the

1 See for instance Smith 2000 to read more about the current state of non-market valuation.
2 Human Well-being depends on the basic requirements for a good quality of life including freedom of choice, health, good social relations and security. Well-being may be broadly understood as happiness. Well-being as experienced and perceived by humans is situation-dependent, as it reflects the local geography, culture and environmental circumstances.
3 Passive Use is now used interchangeably with Non-use or Existence Value. Other terms that have been used include preservation value, stewardship value, bequest value, inherent value, intrinsic value, vicarious consumption and intangibles (Carson et al. 1997).
4 For a more comprehensive understanding of this debate see, for instance, Turner 1999.
5 See Randall 1991 and Smith 1993 for a more comprehensive understanding of the total and non-use values discussion.
6 Environmental resources are increasingly recognised as assets providing services that are no longer readily available. Increasing demand to measure their value and incorporate them into legal, political, and economic decisions is a clear sign of what we would expect as their scarcity grows (Smith 2000).

ecosystem to maintain its properties of self-organisation and stability while enduring stress and shock (Turner 1999).

Use Values include 1:

. Direct Use Values: these derive from the actual use of natural resources for commercial or self consumption purposes (e.g. harvesting timber, fishing, collecting herbs and minerals); tourism and recreation2; education and research3; aesthetic, spiritual and cultural ends;
. Indirect Uses: related with the use society makes from ecosystem functions like watershed values (e.g. erosion control, local flood reduction or regulation of stream-flows) or ecological processes (e.g. fixing and cycling nutrients, soil formation, cleaning air and water). It further includes vicarious use value addressing the possibility that an individual may gain satisfaction from pictures, books, or broadcasts of natural ecosystems even when not able to visit such places;
. Option Values: related with individual willingness to pay a premium to ensure future ecosystem availability and usage;
. Quasi-Option Value: refers to individual willingness to pay a premium to ensure more accurate scientific information;
Passive (Non-use Values) include:

. Existence Value: reflects the moral or altruistic satisfaction felt by an individual from knowing that the ecosystem survives, unrelated to current or future use;
. Bequest Value: considers individual willingness to pay a premium to ensure that their heirs will be able to use the ecosystem in the future.
The economic value of an ecosystem thus relates to the TEV. However, this is not an absolute value. Economics provides valuations only in comparative terms. When they say they are valuing an ecosystem, economists are really defining a trade-off between two situations involving a change: e.g. maintenance or non-maintenance of the ecosystem or PA. The economic value of the ecosystem (PA) is the amount an individual would pay or be paid to be as well off with the ecosystem (PA) or without it (Hicks 1939; Kaldor 1939). Thus, economic value is an answer, mostly expressed in monetary terms, to a carefully defined question in which two alternatives are being compared.
The answer (the value) is very dependent on the factors incorporated in that choice: the object of choice and the circumstances of choice (Kopp and Pease 1997). Economics defines objects of choice as any tangible or non-tangible object, process or activity that can be described as allowing choice. The objects of choice are defined by a set of characteristics and attributes that are perceived by individuals but not necessarily by all individuals. In our case, the object of choice is an ecosystem (PA) whose specificity is defined by a set of environmental and ecological attributes to a greater or lesser extent perceived by individual users and passive users. The circumstances of choice describe the

1 See OECD 1999 for a more detailed definition of the different type of uses. See also Daily 1997.
2 See Mendes 1997 for a Travel Cost estimation of the recreation use value of a National Park.
3 One of the more recognised sources of value is that associated to knowledge. According to a report for the U.S. National Academy of Sciences, the basic source of over $60 billion in current market value was obtained from biodiversity (plants and insects). Furthermore, several of the most productive and robust grain species were genetically derived from wild specimens. Currently, most drug and biotechnological companies are well aware of the strategic, commercial and scientific value of ecosystems as vital depositories of biodiversity. See for instance Chichilnisky and Heal 1998 for more details of these deals.

context in which that choice is made (to accept the political option to conserve the ecosystem or alternatively accepting the political option of non-conservation of the ecosystem). It is clearly fundamental to describe to the individual the consequences of his/her choice, specifically in terms of: i) what is foregone by the choice and what is gained; ii) specify the rights of assignment; iii) define the mechanism of choice, that is the manner through which the individual will exercise choice: by voting, through private market transactions or other unspecified behaviours? See also Metrick and Weitzman 2000 for their interesting parabolic perspective of ecosystem and biodiversity in terms of Noah¡¯s The object and circumstances surrounding such choice define its context. In the case of ecosystems or PAs, value depends on the ecosystem/PA location and the level of human presence, the actual or threatened level of degradation as well as the degree to which natural services provided can be substituted by other substitute ecosystems. This substitutability is a highly important concept within economic valuation as objects with significant numbers of close substitutes are not rated as valuable as those with few or even no substitutes. In the case of ecosystems, the degree of substitutability is relative depending on factors including the scale and level of aggregation and the time-scale involved.

For specifying rights of assignment, there are two possible choice situations. Either the individual gives something up to receive the object of choice that will affect his/her utility or well-being or the individual receives something to give up the object of choice that could affect his/her utility or well-being. The former situation corresponds to Willingness to Pay (WTP) and the latter to Willingness to Accept (WTA) and these are the fundamental monetary measures of value in economics.

These welfare measures applied to non-market transacted objects of choice were first proposed by Maler (1971; 1974) as an extension of the standard theory of welfare measurement related to market price changes formulated by Hicks (1943). Maler stated that it was possible to build four measures of individual welfare change associated to choices involving non-market goods. If the object of choice generates an improvement in individual well-being (a rising utility), two situations become possible. Either the individual is WTP an amount to secure that change, termed Compensated Willingness to Pay (WTPC) or he/she is willing to accept a minimum of compensation to forgo it, the Equivalent Willingness to Accept measure (WTAE). If the object of choice generates a deterioration in well-being (a decreasing utility), again two situations are possible. Either the individual is WTP to avoid this situation, termed the Equivalent Willingness to Pay measure (WTPE) or he/she is WTA compensation to tolerate the damages suffered, the Compensated Willingness to Accept measure (WTAC). When economists talk about the value of an ecosystem or PA they are referring to an individual TEV measured by one of these four welfare measures: WTPC/WTAE if the individual faces an improvement of wellbeing; or WTPE/WTAC where the individual faces a deterioration in well-being.

Maler used the following basic model of individual utility to define welfare measures. Let U = (x, q) be the utility function of an individual with preferences for various conventional market commodities and where consumption is denoted by the vector x, and for non-market environmental amenities denoted q. q may be a scalar where related to a single amenity or is a vector where related to several amenities as is the case of q representing the ecosystem one wishes to value. The individual takes q as given which means q is a public good. It is also assumed that preferences represented by the utility

Arc.

function are continuous, non-decreasing and strictly quasi-concave in x1. The individual faces a budget constraint based on their disposable income m, and the prices of market commodities, p. The individual maximisation utility problem of decision is then formalised as:

Min U(x,q)
x*
subject to p xmmin=¢²

The solution of this problem yields a set of ordinary or marshalian demand functions for x denoted , for i = 1, ¡¦, N individual and an indirect utility function denoted . (iixgp,q,m=()(Ux,qp,q.)))))) )()i,mUgp,q,m;q..==..

The dual is an expenditure minimisation model defined by:

ciixi pxsubject to U(x,q) U min=¢²
The solution of the dual yields a set of compensated or Hicksian demand functions for x denoted , and an expenditure function m. Let us now suppose q is going to change ceteris paribus and the individual i will have to choose between the state q(icixhp,q,U=()(iiiep,q,U php,q,U==¢²0(p,q,=()0()(01Vep,q,Uep,q=.0 or the state q1. If he or she chooses q0 the level of utility is given by U ; if he or she chooses q0m1,U1 the utility is given by U . The welfare change associated to the utility level change can be measured using Maler¡¯s Compensation Variation (CV) or Equivalent Variation (EV) measures, defined respectively by andE. If U(11 p,q,m=00e(p,q,U)e=.1CVp,q,U11 > U0, CV measures WTPC to secure the change, and EV measures WTAE to forgo it. If U1 < U0, CV measures WTAC to tolerate the damage and EV measures WTPE to avoid that change.

WTP and WTA produce differing monetary information for the same object under valuation. Generally, WTA is greater than WTP. This discrepancy stems from the different welfare measure definitions and contexts of choice. The WTP economic value of an object of choice is constrained by individual wealth and by the existence or non-existence of substitutes2. These constraints are not present in the WTA money measure. Besides income and substitutability effects, in the WTA approach the individual accepts money in exchange for something, either to forgo a benefit or to tolerate damage. And as Kanemann and Tversky (1979) stated there is some evidence that individuals experience ¡°loss aversion¡±. This means individuals may be expected to value a unit of loss more highly than a unit of gain where he or she believes that some right to the current amount of environmental asset exists. WTP will though be equal to WTA where there is no income effect, where 1 The specific form of the utility function will affect the shape of the indifference curves. The shape of the indifference curves indicates the preferences the individual has for x and q. In this case, to say the utility function is quasi-concave is merely for the sake of analytical convenience. To say this is realistic or not is considered by utilitarian theory as an empirical question (Hanemann 1999). 2 See Hanemann 1991 for more comprehensive analysis on the WTP and WTA discrepancy when applied to the valuation of environmental services.

there are perfect substitutes of the object under valuation and where the individual is neutral to losses and gains.

2. Estimating the Economic Value of PA Ecosystems

WTP and WTA are the fundamental monetary measures of value in economics. When an economist sets about measuring the value of market goods and services, he or she uses actual, observed, market-based information. Preferences for private goods are revealed directly when individuals purchase them on the market. As ecosystems or PAs are not market tradable, we thus have to elicit individual preferences directly by use of questionnaires, such as Contingent Valuation (CV). CVs provide the means to estimate natural resource value or loss and is the only current method that produces estimations of passive use values1. Furthermore, ecosystems and PAs are natural assets that produce a series of natural services that are directly and/or indirectly used by families and generate benefits associated with passive use, including option, quasi-option and existence values. For purposes of economic valuation, the ecosystem and its components are considered to be public or quasi-public reproducible natural assets, like structures or equipment. As people experience satisfaction with the existence and services of ecosystems (a reproducible natural asset), the value of that reproducible natural asset is necessarily linked to its generated value flows of services and passive uses. Hence, the value of the ecosystem, according to the asset analogy, is equal to the discounted sum of values of those services and passive use benefit flows. When using this natural asset analogy one must be aware that: i) changes in the ecosystem TEV depend on changes in the quality/quantity of service flows and their passive use values; ii) where there are substitutes for ecosystem services, the TEV responds to changes in that substitute; iii) the TEV is dependent on individual income constraint and methodological factors.

To estimate TEV one has to go through the following steps: i) to estimate WTP/WTA via actual individual responses gathered by the CV method; ii) to estimate TEV, using aggregate values of stated WTP/WTA discounted by a certain rate, in accordance with the natural asset analogy.

The Use of the CV Method and Estimate Reliability

The use of CV methods to estimate the theoretical economic measures of TEV for environmental services has been one of the most fiercely debated issues within environmental economic valuation literature over the last twenty years. The principle issue is the validity and reliability of CV estimates and the inclusion or non-inclusion of passive users as a TEV component2. Detractors argue that respondents provide answers inconsistent with the basic assumptions of utilitarian rational choice; they question the seriousness of CV answers because the results of surveys are not binding. A more extreme position holds that the economic concept of value itself has no link to reality. This has led to the supposition that responses might bias CV value away from theoretical welfare measures, WTP and WTA. Defenders of the CV method acknowledge that early applications suffered from many of the

1 For a detailed description of the Contingent Valuation Method see Mitchell and Carson 1989. For a more detailed description of economic valuation methods see Freeman 1993, or Braden and Kolstad 1991, or Hufschmidt et al. 1983. For examples of valuation methods applied to value ecosystem services see Goulder and Kennedy 1997.
2 See Bateman and Willis 1999 to review the practical, methodological problems of the CV method. See also Jakobsson and Dragun 1996.

problems critics have noted (see Mitchell and Carson 1989). However, recognition is required of how more recent and more comprehensive studies have dealt and continue to deal with those objections.

The key to the evaluation method is that it must be assessed in terms of how closely it represents an accurate measurement of the real value of the ecosystem. The closer the real values are to the estimated, the more accurate the valuation method is. If WTP/WTA were observable there would be no problem. But given they are not, it is then necessary to use other complex criteria and ¡°rules of evidence¡± to assess accuracy. In measurement, accuracy means the reliability and validity of data analysis used for the valuation framework1.

From the economics perspective, reliability is related with the accuracy of aggregate WTP over appropriately defined aggregates of individuals. Economists tolerate certain amounts of unreliability in the estimated WTP, if random errors in measurement remain within tolerable boundaries. Thus, the valuation technique reliability depends on the degree of data "noise". The bias between the CV estimated WTP/WTA and the theoretical WTP/WTA grows where the former tends to systematically diverge from the latter.

The concept validity relates to the CV application process. It involves numerous issues that must be resolved mainly based on individual judgement of the CV implementing entity. Because WTP is not observed, inferences as to validity are based on indirect evidence related both with content validity of CV study design and execution and construct validity dealing with the degree to which the estimated money measure relates to other theoretical measures. To assess the content validity of a CV study involves examining study procedure content. This involves four steps.

Firstly, the researcher defines the scenario that would lead a theoretical consumer to reveal his or her WTP/WTA. More precisely, this is the CV phase where the elements of choice or details of the transaction are presented to the respondent. The transaction must be adequately defined in order to be clearly understood by the participant. The second step is vital to controlling the extent to which participants really understand the proposed transaction as communicated through the scenario defined in the first step. This obliges the introduction of qualitative research procedures to support CV survey design. The third and fourth steps refer to the appropriate statistical and econometric techniques that must be applied to elicit unbiased, higher content validity estimates of WTP/WTA.

Contributions towards improving CV studies, from both the theoretical and empirical perspective, have been drawn from the differing fields of academic social science research . economics, psychology, law and politics. (see Kopp and Pease 1997 for a more comprehensive study of this issue). One of the most important contributions was that of NOAA Panel Report2 (Arrow et al. 1993). By recognising the impossibility of externally validating estimates produced by CV studies, the NOAA Panel recommended researchers adopt an ex ante analysis of the results in place of an ex post analysis by focusing discussion on how to improve the theoretical and empirical quality of studies thereby improving the accuracy of CV valuation by strengthening result reliability.

1 See Mitchell and Carson 1989 for a comprehensive description of these methodological CV problems and their potential effect upon estimates. See also Jakobsson and Dragun 1996 for a comprehensive survey of literature on such issues.
2 The National Oceanic and Atmospheric Administration convened a panel headed by the Nobel laureates Kenneth Arrow and Robert Solow to review the CV method for determining non-use value in assessing natural damages under the US Oil Pollution Act 1990.

The Panel guidelines for the study design phase are set out defined for three aspects of the transaction: the good, the payment, and the valuation context. In the case of ecosystem valuation surveys, respondents need to know about the attributes of the ecosystem, the level of provision of those environmental attributes ¡°with and without intervention¡± and if there are undamaged substitute commodities. The researcher must previously determine which attributes (services) affect the value an individual places on a good (Green and Turner 1999; Fischoff and Furby 1988).

As for payment, the Panel recommended the use of the WTP valuation format and the definition of the ¡°payment vehicle¡± which may include taxes, property taxes, sales taxes, entrance fees, changes in the market prices of goods and services or donations to special funds.

As for the transaction context, it is important to explain the extent of the ¡°market¡± by informing respondents of how and when the environmental change will occur as well as the decision rules in use for such provision (e.g. majority vote, individual payment). Researchers must allow respondents the opportunity not to vote. The Panel recommends a conservative survey design, a referendum style choice, and voting choices followed by open-ended questions asking about reasons for voting one way or another. The Panel¡¯s recommendations on survey design are for the most part almost standard practice except: i) the use of a referendum format in substitution of the open ended question eliciting the maximum WTP; ii) and the opportunity to the respondent to choose not to vote. These Panel guidelines are standard practice for any high quality survey. It is recommended to use probability sampling, in-person interviewing, to minimise non-responses, to make careful pre-testing and to examine interviewer effects.

By recognising the impossibility of independent verification of CV results, the Panel suggested that besides the survey design and administration guidelines, an alternative test of CV reliability must be drawn from the economic theory of rational choice to monitor the rationality of individual WTP responses. This test is called the scope (embedding) test and requires that the stated survey WTP should be related to the size of the object of choice1. If the WTP response is inadequate to the scope of the object of choice then the findings of the CV survey are unreliable.

The NOAA Panel concluded that the information provided by CV surveys, where in full compliance with the recommendations, can be considered ¡°as reliable by the standards that seem to be implicit in similar contexts, like market analysis for new and innovative products, and the assessment of other damages normally allowed in court proceedings¡± (Arrow et al. 1993).

More recently there has been a trend to include expertise from other disciplines such as marketing research, survey research, psychology (both cognitive and social) to improve the CV methodology both from the theoretical and empirical point of views. The importance of this contribution to survey research is almost intuitive because CV is broadly survey valuation method based. The contribution from psychology is explained by the fact psychologists consider the model of human behaviour that supports individual conventional utilitarian economic behaviour static and too simplistic. Psychologists criticise the utilitarian approach because individuals do not chose in isolation but are affected instead by the characteristics of their particular social.economic group. They do not choose based on only

1 The scope test is based on the weakest form of rationality among individual choices. It is reasonable to suppose that more of a good is always better to the individual if not satiated and that he/she is willing to pay more for more of that good. Also, it is reasonable to assume that WTP will decline although not abruptly for additional amounts of the good (Arrow et al. 1993).

one restriction such as income. Preferences are not static and they are not equal across all individuals. Psychologists criticise utilitarian economics for the assumption that all values are commensurable and ultimately reducible to a single metric be it money or another type1. Psychologists defend CV responses are sensitive to these methodological factors that in standard economic theory are deemed irrelevant. Underlying much of this discussion are implicit assumptions of what Fischoff (Fischoff 1991) called the philosophy of basic values. The traditional treatment in CV surveys is to consider these responses as protest responses that include: protest bids (zero to infinite bids); non-response; or ¡°unreasonable sacrifices¡±2. When this situation happens one concludes the individual has lexicographic preferences which means he or she bases his or her responses from a hierarchy of values whose structure is dependent upon the strength of individual attitudes, beliefs or dispositions he or she holds towards the valuation context. If there is an excessive number of protest responses the scope and quality of the information for economic assessment of values may be limited because lexicographic preferences violate the assumptions of continuously defined, differentiable and convex preferences in standard utilitarian theory (Rosenberger et al. 2003). Current empirical evidence has been proving such preferences are very common for environmental goods (see Spash 2000 for a review). The psychological dimension to environmental value assessment is an ongoing field and lies beyond the scope of this paper.

Estimating TEV Using Individual Stated WTP

In estimating an ecosystem TEV generating social services, a sample of people drawn from a population has to be asked about his or her WTP for preventing the ecosystem from destruction over a relevant period of time. The type of population one has to consider depends on ecosystem characteristics in terms of size and importance of the services and natural goods provided. Where the CV method has been applied, the answers can be interpreted as reliable and valid estimations of the true individual WTP to maintain the ecosystem. The following question to be discussed is how can we achieve a monetary measure of the ecosystem value from the N individual stated WTP by the CV questionnaire. One approach mentioned earlier is based on the capital asset pricing model.

This approach views the ecosystem value as an asset value (natural and reproducible) at a particular point in time as the discounted value of all future services the ecosystem will provide. A common economic approach is to assume a rate of discount and further assume that the flows of services provided by the ecosystem will be constant and that the value of flows will increase in line with the expected rate of inflation. Under these assumptions, one is left with the task of valuing the services at some point in time discounted at some discount rate. The problem here is what rate of discount is to be chosen and what flows are to be considered.

The answers to either question are not straightforward to economists. This is because the valuation of ecosystems or the decisions involving the conservation or non-conservation of nature, as well as other related sustainability decisions within a more general backdrop, are characterised by several dimensions in which issues are more demanding for economists than those raised by environmental

1 See Green and Tunstall 1999 for a comprehensive introduction to the psychological perspective on economic valuation.
2 In the context of neo-classical definition, ¡°unreasonable¡± may include WTP values far beyond the individual¡¯s ability to pay or sacrifices that involve excessive opportunity costs such as loss of human life.

economics. One is the time dimension. The long-term period applied to ecosystems is much longer than normally considered in economic analysis: not inferior to 50 years and sometimes as long as one, two or even several centuries. This poses a particular challenge for the economist¡¯s traditional practice of discounting. A second dimension is related with uncertainty. Over such a long time scale, it is logical to expect that individual preferences will change due to technical and social changes that also affect the flow of individual WTP across time. On the other hand, the demand for ecosystem services is already so great that trade-offs among services have become a rule. A country can build more roads or construct in PAs and in the process destroy ecosystems for example but in so doing that country is decreasing the supply of natural services that may be of equal or even greater importance. And there are now many indications that human demands on ecosystems will grow still further in coming decades. A formidable increase in demand for and consumption of natural resources, current estimates point to the future existence of a further 3 billion people and the quadrupling of the world economy by 2050, is to be expected (Millenium Ecosystem Assessment 2003). This increase is compounded by the increasingly serious degradation of ecosystem capacities to provide services, seriously diminishing the prospects for sustainable human development (Batabyal et al. 2003). However, it is important to incorporate the monetary measure of the ecosystem TEV into this pressing trend for scarcity. As one cannot ask people to be futurists as to their future WTP patterns, one way of achieving this is to use appropriate discount rates and discount methods that allow for incoorporating growing future ecosystem scarcity. The default criterion that has been used for ranking environmental conservation projects is provided for by the discounted utilitarian approach introduced by Bentham in the nineteenth century. According to this, one must choose the greatest present discounted value of net benefits. The project must be rejected where it provides a negative net benefit. Discounted utilitarianism has dominated, more due a lack of convincing alternatives rather than any intrinsic accuracy.

A positive utility discount rate reflects what economists refer to as a positive time preference1, a widespread desire to consume today rather than save for the future. This is a way by which the market penalises investments with long-term payoffs: any investment with high up-front costs and a long stream of future benefits will dramatically undercount future benefits. This is precisely what happens when one has to value ecosystems by discounting the flow of the services they provide over time: at any positive discount rate, the present value of any ecosystem is almost irrelevant and it thus becomes irrational to be concerned about extinction or conservation. And yet societies obviously are worried about such issues and actively continue to consider how to devote substantial and scarce financial resources to them.

How important then is our concern about the time dimension to ecosystem valuation frameworks? The opinion of some authors is that discounting is ethically indefensible2. However, there is empirical evidence suggesting the legitimacy of discounting future utilities even where it is not certain that discounting future utilities in the evaluation development programmes is the same as discounting benefits of values (Heal 1993). Where empirically proven that people consider some positive discount

1 See also Baumol 1968.
2 This is the opinion of Ramsey 1928 and Harrod 1948, the original proponents of modern dynamic economics and that of Heal 1993 on environmental subjects. See Heal 2000, chap. V for a comprehensive theoretical approach to this issue.

rate we have to conclude that the traditional discounted utilitarianism approach to value benefits is not particularly suitable to clearly (as far as possible) capturing individual future concerns over the future environment. This is because decision-makers and cost-benefit applying researchers generally use market rates of discount much higher than the appropriate efficient sustainable rate1 thus depreciating future flows. As a consequence of this controversy, researchers have, in some cases, begun to apply lower discount rates to long-term, intergenerational projects (see Bazerlon and Smetters 1999). Others use a declining rate in the future. Unfortunately, both methods result in time-inconsistency problems as long-term projects in the present become near-term projects in the future (see Heal 2000 for a more comprehensive discussion of this issue).

More recently, attempts have been made to include the uncertainty and its persistence dimensions into the discount rate discussion that seems to dramatically increase the expected net present value of future payoffs (Newell and Pizer 2003). We may conclude that in any calculation of TEV, we have to choose the appropriate relevant period of time, discount rate and method of discounting in order to reflect intergenerational preferences and uncertainty into the valuation process.

3. So! Just What is Ecosystem Economic Valuation Useful For?

A clear answer to this question suggests it be broken down into other two questions. Firstly, given the existence of such controversy towards theoretical, methodological, and empirical aspects of ecosystem TEV, what is the usefulness of a monetary measure for ecosystem conservation decisions? Secondly, where the economic value is a theoretical, abstract measure experiencing a somewhat exacerbated controversy, what is the point of estimating it and using it for conservation issues? Let us begin with the first.

Many economists and non-economists consider economic valuation of ecosystems as an incentive measure and support for decision-making and to ensure that the private profitability gap between sustainable and unsustainable use of ecosystem services is narrowed or even closed (OECD 1999). Economic evaluation is a support for political and judicial decision-making in a context of accruing environmental degradation especially when compounded by increasing social demand for environmental services. Cost-benefit analysis has often been under-utilised for environmental purposes because the value of ecosystems is both difficult and expensive to quantify although strong progress has been made in the last two decades in the USA and Europe (see Bonnieux and Rainelli 1999, Loomis 1999, OECD 1999, for an overview of the institutional evolution of cost-benefit evaluation techniques).

Ecosystem valuation can also play a beneficial role in government land use, conservation, and tax planning. For instance, there is a growing interest in incentive or compelled conservation by private owners and property developers through economic instruments (Boyd and Winger 2003; see OECD 1999 for a comprehensive description of economic incentive instruments applied to nature conservation). Such economic instruments include tax code changes or the outright purchase of

1 The market rate of discount is associated with income generated by the better alternative economic option foregone and by the fact of the conservation option being chosen. Externalities, non-use, option values, and depreciation of natural capital are not considered. To ecologists the efficient sustainable rate is equal to the growth rate of Net National Welfare (NNW) where NNW = GDP + Normal market output . External Costs . Pollution Abatement Costs . Depreciation of Created Capital . Depreciation of Natural Capital (Goodstein 1999).

properties and easements. Tax code changes and exemptions may be expected to motivate land or easement donations to private charitable organisations and trade development rights where development is geographically restricted and developers have to bid among themselves to compete for the ¡°right to build¡±. Another means used by some governments in pursuing conservation is through outright purchase of properties and easements (Boyd and Winger 2003). All these economic conservation instruments based on cash payments or tax breaks require the estimation of enough compensation to offset habitat losses elsewhere. Evaluation will be particularly important whenever ecological assets are traded. Following the relative success of the air emission permits market, market trading of incentive instruments (tradable development rights for instance) has been proposed as resulting in greater environmental protection. However, in order to guarantee that trade preserves the social value of ecosystems, they require conscientious economic analysis. Badly regulated ecosystem trade may undermine, rather than benefit, the environmental welfare objectives because unlike air emission trading, ecosystem asset trading requires trade-specific environmental values. As we have already concluded, the ecosystem TEV depends crucially on its location, its relationship with human activities and changes over time. Economic valuation methods allow a comparison of different ecosystems, with different values bringing confidence to an ecosystem trading scheme in terms of the maximisation of net social benefits (McGartland and Oates 1995).

There is no consensus about the relevance of using ecosystem valuation as an incentive tool in conservation policies. Some economists argue that evaluation is neither necessary nor sufficient for conservation and defend the key theme of conservation as making this policy more attractive than any alternative usage through translating the social benefits of ecosystems into income (Heal 2000). In their opinion, this does not necessarily oblige evaluation. However, one can question this opinion by asking how one can translate those ecosystem social benefits into income in the absence of markets, and within a context of growing financial resource scarcity? Is valuation not the most reliable, socially just and technically accurate form of answering such a question?

Economic valuation further constitutes a useful tool for educating and involving local populations and stakeholders by highlighting the connection between the ecosystem¡¯s underlying biophysical properties and benefits associated with the active or passive use of its services. It enables the justification of conservation projects encouraging local inhabitants and stakeholders to accept and to comply. If local people are conveniently alerted to the true economic value of their land, including scarcity, they will anticipate a higher future value and hold back from economic activities that may not be compatible with conservation. This behaviour will be smoothly endogenised by local inhabitants and stakeholders where accompanied by decisions that transform conservation decisions into sustainable income for local communities. In sum, environmental valuation can be seen as an economic conservation instrument that promotes ethics and fairness in conservation policies. Through environmental valuation and improved economic compensation mechanisms, society will reinforce its right to defend social ecosystem services even when such a defence imposes severe economic and social restrictions on the use of land that belongs to other people independent of their social-economic development expectations.

Being such an important instrument for convincing people to comply and participate in conservation policy, there remains ongoing debate and a persistent reluctance as to the usefulness, reliability, and validity of the monetary measures produced by economists. Philosophers, ecologists and even some economists not subscribing to the idea of using environmental economic values as conservation tools have their own ethical and philosophical, technical and methodological arguments. It is unquestionable that economic money measures are abstract, theoretical measures and they do not measure the real value as commonly understood. Economists tend to structure their thinking around models of perfect rational agents, with fixed preferences, making decisions in order to maximise certain well defined individual objectives. This is very different from the paradigms characterising the natural sciences or even other social sciences. As a consequence of this hyper-abstract economic model of thinking, economic value is a relative not an absolute measure. It is the answer to a question involving a choice. It indicates how much an individual is willing to pay (to accept) for a unit more (or to forego) of an environmental resource without changing his or her current (future) wellbeing. To say one thing has greater economic value than another is an alternative way of saying that, under the circumstances, this would be chosen in preference to that. A correct understanding of what the economic value concept really means helps in understanding the value paradox1: articles that command great prices are often things that common people consider as being of little intrinsic or useful value, take diamonds or paintings for example. On the other hand, absolutely essential goods are often available at a negligible, or even no price, like water or landscape, at least in regions where such natural resources are abundant. There is, however, no inconsistency where goods that are, overall, immeasurably useful being worth less than the non-essential as economists determine marginal values. A good is worth more than another when the individual is not yet satiated and it is more difficult to obtain an additional unit than another unit of something else. People generally consider a diamond much more valuable than a litre of water, even while knowing the latter¡¯s vital importance, because they recognise the rarity of diamonds and the relative abundance of water. In short, the economic money measure of value basically reflects the scarcity of the good being valued under certain circumstances but not value from a common sense perspective. The economic money measure of value also reflects important issues that affect scarcity as perceived by the individual such as the existence of substitutes as well as the context surrounding the object of choice. In short, one may say that the theoretical, abstract weaknesses of economic valuation, identified by some commentators, is simultaneously its strength when used to improve decisions involving scarce non-market transacted resources, like ecosystems.

Thus, while the level of abstraction and technical rigour of the theory underlying the definition of economic value are seen by some as handicaps, others consider them as trump-cards when there is a need for credible support to political and judicial decisions. However, these are complex methodological problems and the high level of economic and econometric expertise needed to estimate ecosystem values lies at the source of another set of criticisms. It is also a very expensive process. However, these technical, methodological or budgetary reasons must not be used as arguments to turn our backs on ecosystem valuation. They must of course be taken into account

1 The paradox value is one of the arguments used by psychologists to demonstrate the incoherence of the economic monetary

during the political decision-making process when conservation decisions involve very important ecosystems at the local, national, or regional levels. measure and its valuation shortcomings. In short one must not agree with the argument any number is better than no number, used by some proponents, as it denotes an indefensible, very resigned attitude. When talking about economic values, economists are not referring to any number but to a number rigorously and theoretically defined and carefully applied in order to capture the real value people put on the object of choice under certain circumstances of choice. Rather than saying any number is better than no number is to say an economic number is better as a minimum reference number, than any number at all.

4. Conclusions

Traditional biological conservation management practice based on protected areas . PAs . has proved unable to achieve the required conservation levels for PAs with residents, located near population centres and with good access, especially within a low or middle income country context. One of the reasons for this relative failure is that biodiversity in general and PAs in particular do not usually exist in isolation and independent of human activities. For local communities, the act of conservation becomes a misfortune rather than an opportunity for sustainable development leading to increasing evasion of conservation regulations while governments delay in coming up with answers to the growing conservation problems resulting from such individual actions. This is the reason we believe conservation practice has to change deeply. It must not rely only on the SMS principle and command-and-control instruments but must instead engage local residents and other users thus creating a broad consensus over the existence and objectives of conservation initiatives. One way of achieving this is to make people adopt profit-based conservation practices. This is by no means easy for policy-makers being replete with social, scientific and practical difficulties and ambiguities.

In this paper, we have discussed the advantages and disadvantages of ecosystem economic valuation as an incentive measure for enforcing local community co-operation in conservation decisions and management. The utilitarian approach allows value to arise in a number of ways depending on individual use of ecosystems. Hence, economists have generally settled for taxonomy of total ecosystem value interpreted as Total Economic Value (TEV) that distinguishes between Direct Use Values and Passive (Non-use) Values. TEV is a relative value and an answer mostly expressed in monetary terms to a carefully defined question in which two alternatives are compared. This answer depends on elements of choice defining the prevailing context. As ecosystems are not purchased on markets, one has to elicit individual preferences directly by the use of questionnaires such as Contingent Valuation (CV) to assess the individual¡¯s WTP and WTA relevant monetary measures. Complex criteria and ¡°rules of evidence¡± such as those suggested by the NOAA¡¯s Panel must be applied to guarantee the reliability and validity of the CV data. To calculate TEV based on individual CV data and the asset analogy, time and uncertainty have to be considered when discounting service flows. One concludes that TEV may be a useful tool as an incentive, a support for decision-making, and as a tool for education and information. The fact of being a very abstract instrument, and very demanding from the theoretical and technical points of view, becomes an advantage. To date, it is still the only existing, carefully defined and applied and somehow reliable way of society knowing how much an ecosystem is worth within a market-based scenario.

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Copyright © 2003 Global Community WebNet Ltd.