Heathens in the chapel? Economics and the conservation of native biodiversity

David J. Pannell

CRC for Plant-Based Management of Dryland Salinity and School of Agricultural and Resource Economics, University of Western Australia, c/- WA Department of Agriculture, 444 Albany Hwy, Albany WA 6330, Australia


Economic issues related to biodiversity are reviewed and discussed. The paper addresses general principles, as well as specific issues of biodiversity protection or enhancement in agricultural regions of Australia. The elements of an economic decision problem are presented, and used to identify information requirements that need to be met in order for problems of biodiversity management to be fully specified and solvable. Needs are highlighted for better definition of biodiversity objectives, and for improved information about cause and effect relationships between interventions and outcomes. The likely cost-effectiveness of investments in different aspects of biodiversity management in Australia is considered. The importance of paying adequate attention to the farm-level economics of proposed changes in land management is emphasised. This is an important influence on farmers’ responses, particularly if large-scale changes are sought. The use of non-market valuation studies to place monetary values on biodiversity outcomes is discussed. A measured approach to the use of these techniques seems warranted. The use of economic policy instruments for biodiversity is reviewed, and there is a discussion of the question, who should pay for biodiversity interventions? It is concluded that the selection of policy approaches and policy instruments for biodiversity needs to be sophisticated, based on science, and sensitive to different biodiversity-related problems and opportunities.


This paper is a selective review of ideas and tools from economics that can contribute to the definition and achievement of community objectives for biodiversity, meaning those objectives that are adopted through the political process. In discussing the application of these ideas and tools, there is a particular focus on agricultural regions of Australia. My objective is to contribute to the design and selection of biodiversity research to support public programs and strategies. I will emphasise some economic issues that I believe deserve more attention than they often receive from biodiversity researchers, and point out limits to some ideas that have perhaps been too warmly embraced by some (including some economists). There have been a number of other reviews and discussion of economics and biodiversity with different emphases, including Gowdy and McDaniel (1995), Young et al. (1996), Edwards and Abivardi (1998) and van Kooten (1998).

The title of the paper is suggestive of the discomfort that some feel at the introduction of the profanities of economics into the sacred realm of nature. I hope to convince readers, if not that economics can provide a channel to the divine, then at least that it is better to have the devil as an ally than as an enemy.

An over-riding objective is to foster cross-disciplinary communication and understanding. Mascia et al. (2003, p. 650) emphasised the importance for conservation efforts of embracing social sciences:

“Biodiversity conservation is a human endeavor: initiated by humans, designed by humans, and intended to modify human behavior to achieve a socially desired objective—conservation of species, habitats, and ecosystems. Embracing this fact, and recognizing its implications for the nature and use of science in conservation, represents a challenge for academics and practitioners alike. We must all be willing to leave our comfort zone behind, to speak different languages, work in different circles, and accept different beliefs. Communication, collaboration, learning, and mutual respect represent the path to success.”

The potential contributions of economics to biodiversity management vary from case to case. In this paper, I distinguish between cases where the aim is to protect existing biodiversity and cases where the aim is to enhance biodiversity (e.g. by establishing trees to provide new habitat). I also distinguish between salinity-related threats and opportunities, and non-salinity-related issues. Salinity is a particularly prominent environmental issue in Australia (Pannell, 2001a; National Land and Water Resources Audit, 2001) and its predicted consequences include major impacts on biodiversity (George et al. 1999b). The explicit consideration of salinity reflects the paper’s genesis – it was prepared for a workshop on biodiversity research priorities for the Cooperative Research Centre for Plant-Based Management of Dryland Salinity.

The paper begins with discussion of an economics-style decision framework and its implications for research into and decision making about environmental interventions. There follows a brief outline of the types of interventions that are relevant to the different categories of problems and opportunities being considered here (protection vs enhancement, salinity vs non-salinity). Then I will present and discuss a range of economic issues at the farm level. This leads to discussion of issues surrounding non-market valuation of biodiversity (placement of monetary values on relatively intangible benefits) and the use of economic policy instruments to protect or enhance biodiversity. Finally implications for researchers and for policy makers are identified and discussed.

As long as it is, the paper does not address all issues that are relevant to the application of economics to biodiversity conservation. Those included are selected as being particularly relevant and/or currently topical in Australia. A list of some issues that are not dealt with is included before the conclusion.


Economics is largely concerned with decision making, particularly with decisions where scarce resources have to be allocated among competing uses and where, as a consequence, trade-offs are required. Economists could therefore be helpful to environmental managers, many of whom are interested in questions that correspond exactly with this type of problem. One example is the question of how to allocate limited funds among many alternative projects to protect environmental assets.

There is considerable scope to improve the efficiency and cost-effectiveness of allocation decisions in some Government environmental programs in Australia (Pannell 2002). Possingham (2001) and PMSEIC (2002) argue that past efforts to address biodiversity problems have not been efficient because, often, the problem is not properly posed, objectives are not clearly stated, constraints are not identified and relevant theory and data are not used in decisions.

To illustrate the potential gains from a more thorough and rigorous approach, consider the BushTender trial scheme in Victoria. It was found in the trial that an improved allocation process for funding to protect remnant vegetation, with greater attention to maximising outcomes per dollar spent, can dramatically increase the resulting environmental benefits (Stoneham et al. 2002). Compared with the approach used in typical fixed-price subsidy schemes to encourage fencing of remnant vegetation, and using a particular system for rating the environmental outcomes, the BushTender approach improved the cost-effectiveness of expenditure by between 7 fold and 30 fold (Gary Stoneham, pers. comm. 2003).

One general contribution that economists could make here is to clarify the elements of allocation decision problems, and thereby detail the information requirements for the problem to be fully specified and solvable. Put most simply, the elements consist of one or more objectives, some alternatives or options, some constraints that must not be violated, and some technical relationships describing how different options (or combinations of options) will affect the objective(s) (i.e. response functions that describe cause and effect). Possingham (2001) has argued cogently for the use of this “decision theory” approach to decisions about biodiversity priorities and discussed its application to several types of biodiversity-related issues (fire management, reserve system design, threatened species funding). PMSEIC (2002) applied the approach to identify broad priority policy actions for biodiversity.

Possingham (2001) list seven stages in the application of a decision theory approach to biodiversity issues.

1. Specify the management objective

2. List the management options (the decision variables)

3. Specify the current state of the system

4. Develop a model of the dynamics of the system being managed.

5. Specify constraints that limit the decision variables

6. Be honest about we don’t know (specify ranges of uncertainty)

7. Find solutions to the problem

To these I would add:

8. Investigate the sensitivity of the solutions to changes that reflect our uncertainty about the parameters of the problem.

I suggest that two of these stages, 1 and 4, pose particular problems for biodiversity. They are discussed a little further here.


For financial problems, selecting the objectives is often easy (e.g. maximum profit, or a balance between profit and risk), but this is not the case for biodiversity problems. Maximising the level of “biodiversity” per dollar spent would seem a potential candidate for an objective of environmental policy. However, definitions of biodiversity tend to be broad and non-specific (Zeide 2001). An extreme example comes from Wilson (1994, p.359) who defines biodiversity as:

“the totality of hereditary variation in life forms, across all levels of biological organization, from genes and chromosomes within individual species to the array of species themselves and finally, at the highest level, the living communities of ecosystems such as forests and lakes.”

A similarly broad definition is given in the National Strategy for the Conservation of Australia’s Biological Diversity.

“the variety of all life forms – the different plants, animals and micro-organisms, the genes they contain, and the ecosystems of which they form a part.” (quoted in Young et al. 1996).

These are not helpful definitions from the perspective of providing an objective for public policy or private decision making. One could not use them to select one environmental project ahead of another. Practical decision making requires much more specific objectives to be specified. A key task for ecologists and environmental scientists is to provide advice to the community on specifically which environmental outcomes are more and less important, including supporting rationales for this advice.

One possible response to this challenge is to say that we should find out which environmental outcomes the community wants and values, perhaps by conducting non-market valuation studies of the alternative outcomes (See Loomis and White 1996; Nunes and van den Bergh 2001; Bräuer 2003 for support for such a view. See Gowdy 1997 for reservations about it). Such a response sidesteps the issue. Non-market valuation studies may indeed be useful, but the quality of the information they provide is constrained by the quality of information we have about the ecological, environmental and social significance of the environmental assets being valued. Collection and interpretation of such information is a task for ecologists (among others) and it is a task that has barely commenced (at least in terms of providing information that is useful for community decision making). The significance of particular species or habitats to the community may depend on biology-related characteristics such as

as well as on socially-relevant characteristics such as:

Biological scientists might, perhaps, feel that the former list is a more respectable target for their attentions than the latter, but the practical reality is that the items in the latter list are relevant to the community. Consider, for example, whether the arguably-more-frivolous factors on the list would influence the results of community surveys, including non-market valuation studies.

This is not to say that it is ecologists’ role to find out what is important to the community in the way of biodiversity conservation. As I am conceiving it, ecologists (in concert with other specialists) would provide information that helps the community (and/or its representatives and institutions) select specific objectives. The objectives chosen should be specific, measurable and supported by a rationale. The objectives should not fully specify the courses of action to be taken (since that pre-empts the decision and may rule out more cost-effective options) but rather should specify the desirable outcomes.

Ideally, the objectives would be related to and expressed in terms of the high-level benefits to society that biodiversity provides (e.g. specific ecosystem services, a source of genetic resources for medicine or agriculture, values related to recreation and tourism, non-use values such as existence value). In practice, planning on the basis of such broad high-level objectives is very difficult, primarily because of lack of information (e.g. about the contributions of different biodiversity outcomes to each of these objectives).

A practical fallback position is to select narrower and more specific objectives for biodiversity outcomes that are judged to provide high levels of one or more of the relevant values. A good example is provided by “Victoria’s Native Vegetation Management: A Framework for Action” (Department of Natural Resources and Environment 2002), which selects an objective of overall “Net Gain” in vegetation values based on a specific vegetation quality rating system known as “habitat hectares”. Some other Australian policy documents express their objectives in broader, non-specific terms, such as “Protect and restore high value wetlands and natural vegetation, and maintain natural (biological and physical) diversity”.

Benefits and costs

Overlaid on the biodiversity objectives discussed above, economists would generally subscribe to a constraint that the benefits of action (broadly defined) should exceed the costs. This is not to advocate for a full benefit:cost analysis for every possible investment, which would be both problematic and expensive for biodiversity (see the section on “non-market valuation of biodiversity” later). Rather it should be seen as a need for investment in biodiversity to be balanced against other possible investments. It is not about economising for economy’s sake. It is a recognition that the “opportunity cost” of public money is important to the community. Funds spent on biodiversity are not available to be spent on other important public services or amenities (health, education, infrastructure, welfare, the arts).

Furthermore, expenditure of any public funds has the opportunity cost that, as a consequent of their removal from private individuals or businesses through the tax system, the funds will not be spent in the ways that would have been most beneficial to those people, as judged by them. The positive-net-benefit criterion of economists, which is sometimes interpreted as hard-hearted or even anti-social, is actually a reflection of soft-hearted and pro-social objectives on a broader canvas.

Economists particularly decry calls for an outcome to be pursued at any cost. At any cost is too high. Unfortunately, discussions of biodiversity are sometimes couched in terms that imply these extremes.

Response functions

For any given biodiversity objective, what are the cause-and-effect relationships describing how outcomes will vary for different types and intensities of intervention? Although we can make reasonable predictions in some cases, in many others the quality of available information is not high. In part, no doubt, this is because it is intrinsically a very difficult task. Ecosystems are complex and there is considerable uncertainty about how they may respond to changes in management.

Consider a question such as, “If I protect 25 or 50 percent of a reserve that is predicted to be entirely lost to some form of degradation, by what percentages do I improve the probability of survival of a particular threatened species that occupies the reserve?” This sort of question will never be easy to answer. Nevertheless, it is a question of the type that scientists need to help the community address if we are to make sound choices about the allocation of public funds for environmental programs. I suspect that it is also important to be able to answer such questions if environmentalists are to convince the broader community that greater levels of public investment in the environment are justified and can be spent well. As van Kooten (1998) observes, a “challenge of biodiversity research is to make a stronger case for sacrificing economic development benefits for increased protection of biodiversity,” (p. 13).


As noted earlier, I distinguish threats and opportunities for biodiversity on two criteria: salinity-related versus non-salinity-related, and protection of existing habitat versus creation of new habitat. The relevant interventions vary among the four categories defined by combination of these two criteria. Table 1 is a summary of the relevant interventions. Subsequent sections will refer back to these four categories.

Table 1. Examples of interventions to promote biodiversity.

  Salinity related Not-salinity related
Protection/ enhancement of existing habitat Implementation of strategies (plant-based or engineering) to protect remnant vegetation, wetlands or aquatic biota from impacts of rising saline groundwaters. Protection of remnant vegetation from clearing.

Fencing of remnant vegetation to prevent grazing by livestock.

Creation of new habitat Adjustment to plant-based salinity management systems in order to enhance their direct values as habitat. Establishment of non-commercial trees for enhancing biodiversity.



This section provides observations and inferences on the likely overall cost effectiveness of public investments in biodiversity in the four categories. Although one cannot speak definitively about the economics of such broad categories, they have sufficiently distinct characteristics for some broad generalisations to be reasonable. In all cases I am referring to outcomes over the long term, not just in a single year.

Protecting existing habitat from salinity

Recent research and modelling by hydrologists has led to a dramatic re-assessment of what it takes to prevent impacts from rising saline groundwaters. Whereas it was previously hoped that small areas of cleverly placed perennial vegetation may be sufficient, the new consensus is that in most regions, very large areas of perennials are usually required to make a worthwhile difference to even relatively small-scale salinity impacts (e.g. Hatton and Salama 1999; Hatton and Nulsen 1999; George et al. 1999a; Campbell et al. 2000; Stauffacher et al. 2000).

To illustrate the point with one example, Stauffacher et al. (2000) evaluated strategies for revegetation with deep-rooted perennials on well over 50 percent of the Wanilla catchment in the Eyre Peninsula of South Australia. They found that this scale of revegetation would be necessary to reduce by around 5 or 6 percentage points the area of land predicted to be moderately-to-severely salt-affected by 2020. Other analyses have found that this low level of responsiveness to revegetation is common in low-to-medium rainfall farming regions.

This new perspective on revegetation has profound implications for the cost-effectiveness of revegetation with perennials as a strategy to protect biodiversity from salinity. It makes the practicality of the approach largely dependent on whether the perennials are commercially competitive with traditional agricultural practices. If available perennial species are not commercially competitive on the required scale (which is currently the case for most of the low-to-medium rainfall zone – Kingwell et al. 2003) the cost of the revegetation would be economically crippling to a farmer. If borne by a government, it would be a major drain on environmental funding.

It seems that direct public investment to fund establishment of perennials specifically to protect threatened biodiversity from salinity is only likely to be a cost-effective strategy in special cases. It may be justified where all or most of the following conditions were met: exceptionally high environmental values at risk, high levels of threat to those values, responsive (localised) groundwater flow systems, perennials available with profitability that is not too far below traditional agricultural land uses, and a location that is close to other valuable assets that would also benefit from the revegetation.

The strategy could be applicable to outstandingly important pieces of remnant vegetation, probably larger areas on public land, or to wetlands that happen to be located where they can benefit when revegetation is implemented to protect water resources (i.e. where it generates multiple benefits). From an environmental or an economic perspective, direct public investment in small remnants on private land would not usually stack up as a strategy for protection from salinity.

It is worth acknowledging that in Western Australia, the state worst affected by dryland salinity, there is a large and diverse flora and fauna evolved for adaptation to saline conditions. An issue for this biodiversity is not protection “from salinity” as such but protection from salinity-related and non-salinity-related changes in the habitat. The salinity-related changes include increased risk and duration of flooding, increased waterlogging and, in some cases, increased concentration of salts beyond tolerated levels (e.g. concentration at the soil surface through capillary action and evaporation).

Protecting/enhancing existing habitat facing other threats

In this category, the threats to biodiversity may include processes such as clearing, grazing or weed invasion. Assuming that these threats can be mitigated using relatively localised and inexpensive measures, such as fencing to control grazing, the cost-effectiveness of investing in this category is much more likely to be high than for the previous category. In other words, a strategy that maximised environmental benefits from a fixed budget would probably favour this category over the last, except in the special cases I outlined above.

Compensation to farmers for giving up rights to clear native vegetation has been publicly advocated by some. A relevant question is whether compensation should be offered, or whether farmers should be required to bear the cost. This is related to whether the farmers have, or are perceived to have, the right to clear. The level of compensation would presumably be based on the “opportunity cost” of the land (the income sacrificed by not using it for traditional agricultural uses), minus the costs of clearing and developing the land.

Given the variation in the extent of the right to clear in different states, it is apparently a matter of community attitudes, and these may change over time. The general question of who should pay is discussed more fully later. My simple analysis suggests that, other things being equal, such compensation is likely to be more cost-effective than supporting revegetation to manage saline groundwaters that are threatening native vegetation. I am not arguing that farmers should necessarily have the right to compensation (see “who should pay?” section for more discussion of this question), merely that such compensation may be more cost-effective at providing biodiversity than another investment option.

The strength of this conclusion would depend to some extent on the biodiversity value of the large-scale salinity-oriented revegetation. (As that value increases, the conclusion in favour of clearing compensation would be softened to some extent.) I anticipate that this will vary widely depending on the spatial layout and species choices, with commercial considerations tending to compromise biodiversity values in many cases. At best, the biodiversity values of salinity-oriented revegetation would approach those of special purpose biodiversity plantings, which are considered in the next section.

Establishing new vegetation primarily for biodiversity

The Landcare program has encouraged and the Natural Heritage Trust has partly paid for farmers to revegetate with non-commercial species of native trees, specifically for biodiversity purposes. The overall cost of such revegetation is the opportunity cost of the land for agricultural uses, plus the direct input costs of revegetation. For similar land, the cost of this strategy would be higher than that for preventing clearing (which, recall, is opportunity cost minus the cost of clearing). Further, the biodiversity value of revegetated land, even if designed with biodiversity in mind, would presumably be less than that of native vegetation that has never been cleared.

On the other hand, other things being equal, this category of revegetation may be more cost effective than revegetation to prevent salinity damage to habitat. This depends on factors that are highly spatially variable and possibly interconnected, such as the biodiversity value of the protected habitat, and the biodiversity value of the salinity-oriented revegetation, among other things.

Exploiting the creation of new habitat from commercial perennial species established

There is a potential opportunity to enhance biodiversity in concert with efforts to revegetate land with perennials in order to prevent a range of salinity impacts, or to make productive use of salt-affected land (even if the revegetation is conducted primarily for reasons other than protection of biodiversity). Notably, if we see success in current efforts to develop woody perennials that are commercially viable on a very large scale, there appear to be prospects for high contributions to biodiversity (e.g. though provision of buffers, linking remnants, and direct provision of habitat).

The key benefit of the commercial orientation now being taken by many Australian agencies, under the leadership of the Cooperative Research Centre for Plant-Based Management of Dryland Salinity, is the promise of much larger scales of woody revegetation than can be achieved through direct public support to offset farmers’ expenses. Interestingly Paine et al. (1996) argue for a similar approach in the USA, with short-rotation woody crops being seen to have the potential for multiple benefits, including carbon storage, reduced soil erosion, improved water quality and quality habitat for wildlife.

At the simplest level, the existence of large areas of woody perennials of various types seems very likely to provide greatly expanded habitat for some species. Whether these species would be conservation priorities is not yet clear. A further possibility is the identification of low-cost refinements to the location, layout, species selection or management of the woody perennials, such that the suitability of the plantings as habitat is enhanced. If such refinements are sufficiently low cost, then this category could provide opportunities for highly cost-effective investments in enhancement of biodiversity.


In the discussion above, landholders are presented as the producers or suppliers of biodiversity benefits. However, one benefit of biodiversity protection is personal satisfaction for the landholder, which is, in a sense, a “consumption” benefit. Michael (2003) proposes that “the literature usually overestimates t he opportunity costs of preservation” (p. 243) because it fails to take account of the willingness of farmers to bear some of the costs, reflecting the personal satisfaction that they gain by contributing to biodiversity protection, or perhaps gains in productivity on other land. This is supported by empirical evidence from the BushTender trial in Victoria, where landholders submitted bids in an auction, specifying the level of financial support they would require in order to undertake specified works to protect remnant vegetation.

“The diversity of bids, particularly the fact that some landholders offered very low bids per hectare, implies that some landholders were probably prepared to share costs with the government to conserve biodiversity. Other landholders, it seems, charged NRE [the Department of Natural Resources and Environment] the full opportunity cost of land-based activities.” (Stoneham et al. 2002).

In the Landcare and Natural Heritage Trust programs, we have observed farmers contributing considerable time and resources to on-ground works intended to promote biodiversity. Under an economic interpretation, one of the objectives of Landcare could be expressed as increasing the farmers demand for (i.e. willingness to pay for or contribute to) environmental improvements. It has succeeded to some extent. One might speculate that farmer’s willingness to pay for biodiversity would vary among the four categories of intervention outlined earlier. Historical experience indicates that at least some farmers object in principle to proposals to reduce clearing, even if compensation is offered, and especially if the proposal involves compulsion. It seems likely that the willingness of farmers to share costs would be lower in this circumstance, reducing to some extent the cost effectiveness of clearing compensation as a biodiversity strategy. The extent of the reduction is an empirical question.

Even in the other categories, we need to have realistic expectations about the extent of farmers willingness to pay. “To a greater or lesser extent, almost all farmers are willing to make financial sacrifices for the good of their land or the environment, but they also must give priority to remaining in business and meeting other family and social objectives.” (Pannell 2001b, p. 131). The point is that the community can benefit from the generosity and environmental concerns of farmers, but that there are limits to what can realistically (or reasonably) be expected.

A common finding in economics is that most consumers of most goods experience “diminishing marginal utility” as their level of consumption increases (e.g. Wonnacott and Wonnacott 1990). This means that the level of satisfaction for each additional unit of consumption falls as consumption reaches higher and higher levels. This is reflected, for example, in the fact that as the price of a good rises, consumers tend to choose to purchase less of the good (demand curves slope downwards). Turning that observation around, to encourage consumers to purchase more of a good (assuming tastes and preferences are fixed) one must reduce the price of the good (Nicholson 1989).

I suggest that the extent to which farmers have given of their time and money for environmental works in the recent past is a good indication of their personal benefits as consumers of environmental benefits (in the sense of deriving personal satisfaction from the results and from contributing to community-wide benefits). Attempts to encourage farmers to expand their contributions will come up against the influence of diminishing marginal utility. Substantial increases in the demand for biodiversity goods by farmers on their land may require substantial falls in the “price” to them of purchasing the good. Two possibilities for achieving this are: provision of more substantial subsidies by the public, and development of less costly technologies for revegetation (e.g. woody perennials that are more commercially attractive and so have lower opportunity costs). This is not to deny that there is scope for increased voluntary contributions from some farmers. Deeper knowledge of environmental issues from participation in government programs can be a strong motivating force for some farmers. In economic terms, this would be expressed as an increase in demand, rather than a decrease in cost of supply.


This section considers a range of issues in evaluating the costs and benefits of biodiversity enhancing actions at the level of an individual farm. Although biodiversity enhancement has benefits (and possibly costs) outside the farm, economic analysis of environmental works at the farm level remains important (Pannell 1998).

Notwithstanding the emphasis in Australian Natural Resource Management on regions, catchments and farmer groups, the individual farm remains the seat of decision making about what actually happens on the farm. Farmer adoption of conservation practices is influenced by many factors (Pannell 1999b), but it is critical to understand the economic incentives they face at the individual farm level (Sinden and King 1990; Cary and Wilkinson 1997).

Analyses at the farm level can also contribute in crucial ways to the more comprehensive analyses. For example, they provide

(a). information about the environmental outcomes that will result if an individual farmer pursues management practices that are optimal from the perspective of financial self interest.

(b). information about the shadow price of land of different soil types in different regions, which indicates the opportunity cost of land converted to other uses for conservation purposes.

(c). information about the trade-offs between profit and environmental objectives at the farm level.

(d). information about the likely impacts of particular policies or incentive systems on farmer behaviour and consequently on environmental outcomes (Pannell 2003), indicating levels of financial support that may be needed to achieve particular environmental outcomes.

See Pannell (1998) for examples.

Identification of farm-level benefits and costs

The elements of a farm-level economic analysis of a biodiversity intervention from Table 1 may include the following.



Some of these elements are relatively simple to identify and estimate in an economic analysis (e.g. direct harvest value, direct input costs) but other are much less straightforward (e.g. opportunity costs, impacts on risk and flexibility). The benefits and costs are affected by a diversity of policy choices on issues such as tenure, infrastructure, R&D, and trade. People without farm-management expertise often do not appreciate the complexity and difficulty of meaningfully quantifying all relevant costs and benefits of changing land use.

To take one example, the opportunity cost of land depends on the best agricultural use for that land, and the interactions (economic, physical and managerial) between that land and the rest of the farm. Table 2 lists a range of factors affecting decisions about possible agricultural land uses.


Table 2. Factors affecting decisions on production of alternative agricultural products

1. Short term profit factors: yields, prices, costs, output quality

2. Dynamic factors (short to medium term): year-to-year carry over of nutrients, weed seeds, pasture seeds, feed, plant disease

3. Sustainability factors: soil degradation (acidification, organic matter decline, erosion, nutrient decline, non-wettability), herbicide resistance

4. Risk factors: yield variability, price variability, yield/price covariance, enterprise flexibility, the farmer's attitude to risk

5. Whole-farm factors: Total crop area, machinery capacity, total feed supply (timing and quality), feed requirements of livestock on hand (timing and quality), finance, labour, the farmer's objectives (profit, risk reduction, sustainability, leisure), the farmer's knowledge and experience

Source: based on Pannell (1995).


Encompassing all of these factors in an economic analysis is a non-trivial task, even before one gets to the subtleties of dealing with biodiversity-related aspects. Economic models are available that bring together combinations of these factors that are appropriate for addressing different types of problems at different scales. The main point here is that one should not take for granted or neglect the farm business aspects of analyses of the economics of biodiversity interventions. The farm-level economics of biodiversity interventions (Table 1) will vary widely from farm to farm (Stoneham et al. 2002; Pannell 1999a), in significant part because of variation in variables related to the farm business rather than to biodiversity per se.

The importance of the economic concept of opportunity cost has been apparent in the discussions above. Another important concept is that of marginal benefits and costs. The issue is that in considering the level of supply or demand for a product, people consider marginal, rather than average, costs and benefits, or at least, they are likely to do so if they are considering the benefits and costs carefully. For example, suppose a farmer is considering planting a commercial woody perennial that also has biodiversity benefits. The farmer may find that if 500 hectares is established, a positive average profit can be made across the 500 hectares, but that if only 400 hectares are planted, the overall net benefits would be greatest. For this farmer, the marginal net benefits of the woody perennial would be positive up to an area of 400 hectares, and negative thereafter. A decision to establish 500 hectares despite this would need to be interpreted as reflecting a laudable private willingness by the farmer to share biodiversity costs, rather than as a solely commercial strategy.

The same logic applies even if the outcomes are not financial. If we look at changing design, location or management of plantings to enhance biodiversity, an economic approach to considering the changes would ask, what is the marginal cost of the change, and what is the marginal benefit (e.g. the resulting increase in biodiversity value, not the total). This implies a fairly sophisticated and sensitive capacity to quantify biodiversity value, not necessarily in financial terms, but at least in some meaningful ecological terms.

The following subsections provide further brief comments on issues in the conduct of farm-level economic analysis in each of the categories of Table 1.

Protecting existing habitat from salinity

As noted earlier, this is likely to be a very high-cost strategy unless the farmer is able to plant perennials that are commercially competitive with the best existing agricultural land use. Therefore, the key economic issue for this category is the economic performance of commercial perennials. There are many studies that have addressed this issue in Australia. In some situations, the available woody perennials are commercially attractive. Most commonly, this occurs in relatively high rainfall regions (Heaney et al. 2000; Stirzaker et al. 2000). In low-rainfall zones, there are currently relatively few commercially proven perennial options apart from lucerne pasture (Kingwell et al. 2003). The area over which lucerne is economically attractive varies depending on factors such as the soil types of the farm, and the market conditions for animal products (Bathgate and Pannell 2002). Typically it is commercial on a scale that would delay, but not prevent, off-site salinity impacts.

Use of engineering methods is also relevant and is being applied, for example, in the case of Lake Toolibin in Western Australia. It too is an expensive approach. I am aware of no current information about the cost effectiveness of using engineering methods to protect biodiversity.

Protecting/enhancing existing habitat facing other threats

If the strategy being considered is compensation for not clearing then there may be a need to assess the opportunity cost of the land, which brings with it a need for the kind of farm-economics analysis discussed above. On the other hand, if the issue is about improved management and protection of vegetation that cannot legitimately be cleared (e.g. because of existing legal constraints that are present in some states of Australia), the relevant farm-level economic analysis would be very much simpler.

Establishing new vegetation primarily for biodiversity

In this category, the opportunity cost of farm land is again a critical part of economic analysis of the strategy. O’Connell (2002) examined the economics of revegetation with non-commercial woody perennials in the eastern wheatbelt of Western Australia. Figure 1 shows the results of one scenario. It includes the annual opportunity costs (around half the total) and the direct establishment costs amortised over 20 years.


Figure 1. Impact of revegetation with non-commercial woody perennials on farm profit in the eastern wheatbelt of Western Australia (Source: O’Connell 2002).


As the figure shows, the context of O’Connell’s analysis is the use of trees to reduce groundwater recharge, but the results give a good indication of the opportunity costs from large-scale revegetation strategies for other purposes as well. The results focus on revegetation of the two sandy soil types on the farm, representing 30 percent of the farm area. Complete revegetation of both soil types would result in a total cost of $90,000 per annum (the vertical distance between lowest and highest points of the graph.

Exploiting the creation of new habitat from commercial perennial species established

It has been suggested that refinements to the location, layout, species selection or management of commercial woody perennials may contribute to their biodiversity value. The required economic assessments of these changes will depend on their nature. If they are simply changes to management practices such as harvesting frequency, a simple analysis may suffice. Changes to the location and scale of revegetation would probably require an analysis that encompasses the opportunity costs of affected land.


There has been increasing interest in economic techniques to ascribe monetary values to non-traded environmental values, such as biodiversity. Environmentalists appear somewhat divided on the matter. Some are supportive of the idea, perhaps hoping to encourage policy makers to give due weight to environmental outcomes alongside more tangible financial outcomes. Others appear to be affronted at the very idea that environmental outcomes could be considered in monetary terms. Interestingly, the techniques are also controversial in economics, over issues that only partly overlap with the concerns of environmentalists.

Methods for valuing biodiversity

Several methods for estimating non-market values have been developed and applied by economists. Smith (1996) outlines three indirect methods (“travel cost recreation demand”, “hedonic property value (or wage) equation”, and “averting behaviour or household production model”) and one direct method (“contingent valuation”). More recently, another indirect method called “choice modelling” (Bennett and Blamey, 2001) has increased in prominence.

The non-market values of biodiversity (particularly non-use values) are usually not measurable using the first three methods mentioned above. Contingent valuation (CV) and choice modelling (CM) are more relevant. Both are based on social surveys of samples of the population. In CV, people are asked to state their willingness to pay for a hypothetical improvement in environmental quality or their willingness to accept compensation for hypothetical deterioration in environmental quality. In CM, people are asked to rank hypothetical options that involve trade-offs between environmental and other outcomes. Values attributable to the environmental outcomes are inferred from their responses. In both cases, values are aggregated up to the level of the whole population.

Debate on the validity and usefulness of non-market valuation methods

There has been a spirited academic debate about the validity and usefulness of non-market valuation methods, particularly for CV. Arguments put forward by advocates of the techniques have included the following:

(a) When done well, the techniques give plausible and realistic results; and

(b) Even though the techniques are not perfect, it is important to attempt to measure non-market values using the best available methods because it assists in having them fully and properly considered in public planning and policy making.

On the other hand, some economists reject argument (a), particularly for CV. For example, Plott (1993), in summarizing the findings of contributors to Hausman (1993) says that:

The basic conclusion of all the papers is that CV should be discarded as a public-policy tool for determining economic damages to the environment. (p. 467) … The conclusion … is expressed at each of four different levels of analysis. (1) The numbers are too variable to be reliable. (2) The numbers do not measure what they are supposed to measure. (3) In fact, the object to be measured by the methods might not be measurable at all. (4) Finally, … the appropriateness of CV for assessing damages, as opposed to more procedural methods, is challenged. (p. 477)

Much of the debate is technical (mostly based on arguments that results from actual CV studies are illogical in a variety of ways). The strength of opinions can be gauged by the following quote from Desvousges et al. (1993, p.114) who conducted tests of the validity and reliability of CV.

At the outset, we believed that it was very difficult to estimate nonuse values accurately by using CV. However, we also thought that it could be done with scrupulous attention to detail, sufficient time and generous funding. After months of listening to conscientious respondents trying to answer difficult questions and of intensively analyzing our data, we cannot maintain our initial confidence in using CV for measuring nonuse values. Given the current state of the art, we do not think that CV provides either valid or reliable estimates of nonuse damages.

CM avoids some of the problems associated with CV and its advocates claim that it is a superior technique. These claims seem reasonable. However there are some more general concerns about CV that would also affect CM.

The first relates to the reliability and validity of survey-based techniques in general. Foddy (1993) discusses cases where it has been possible to check the answers to simple, factual questions asked in surveys, such as “Do you have a driving license?” Typically, error rates in responses to such simple questions are between 5 and 17 per cent. This raises questions about the capacity of surveys to probe more subtle or complex questions, such as non-market environmental values.

Secondly, the great majority of people surveyed have low levels of knowledge of the complex issues about which they are being surveyed. They obtain some knowledge from the survey’s introductory material, but realistically such knowledge would not be deep. Clark et al. (2000) found that survey respondents themselves were concerned about this. Diamond and Hausman (1993, p. 30), in discussing this issue, argued that

It makes no more sense to rely directly on ill-informed members of the public to evaluate the dollar value of such environmental damage than it would be to rely on an ill-informed public to choose between alternative designs for airplanes or nuclear power plants.

With sufficient investment of time, effort and interest, most people would, no doubt, be able to express meaningful opinions on the specific environmental issues being examined, but they are not given this opportunity in a survey.

Argument (b) (that we need to measure non-market values so that they are given due weight in policy and planning) may perhaps have some merit. One potential concern is whether the improvement in decision making is sufficient to justify the considerable expense of conducting valid and reliable surveys, estimated by Dumsday (2003, p.34) to be “anywhere from $40,000 to $200,000”. In a compromise approach that moderates the expense, work is underway to systematise a process of “benefit transfer” (e.g. Morrison et al. 2002) in which a database of past non-market valuation studies is used to provide indicative valuations that are relevant to new issues.

Despite argument (b), some continue to believe that it is adequate to quantify environmental outcomes in terms that are biologically meaningful (such as the number of species affected, or the area of habitat affected, or the increased probability of preventing extinction of a species). This does not avoid the fact that somebody still has to weigh up the information and make a choice between the management or policy options, and that this choice implicitly values the environmental assets at some monetary value. But it does not follow from this observation that a non-market valuation survey is mandated. As one alternative, there is interest in approaches such as citizen’s juries (e.g. Aldred 2001), in which choices are made by lay people who are provided with assistance to access and understand detailed scientific knowledge about the alternatives and their consequences. One advantage of that approach could be that participants have the option of deciding that the science is not sufficient, and that further research is required before a decision can be made. This important option is, of course, not available to respondents to CV or CM surveys.

Finally, I reiterate that, regardless of the process and the techniques used, the quality of the outcomes is limited by our ability to answer basic (non-monetary) questions such as:

It appears to me that research to improve the quality of answers to these question is probably more important than research to value the results in monetary terms.


There has been a dramatic growth of interest in the potential use of economic policy instruments for environmental outcomes (e.g. ABARE 2001; Pannell 2001c) including for biodiversity (e.g. Young et al. 1996; Bowers 1999; Fernandez 1999; Gibbons et al. 2002; PMSEIC 2002; Doremus 2003; Gerowitt et al. 2003), to complement the range of non-economic policy instruments currently in use. Here I present a brief overview of the rationale and scope for their usage for biodiversity.

The rationale for economic policy instruments

Economists recognise two main requirements for the use of economic instruments (or indeed any type of policy intervention) to be desirable. The first is that there should be one of a number of recognised causes of “market failure”. These are cases where allowing the invisible hand of the free market to operate without constraint may result in inferior and inefficient outcomes from the perspective of the broader community (Randall 1981). Potential causes of market failure, which are explained in any resource economics textbook, include:

At least four of these types of market failure are relevant to some aspect of biodiversity (with monopoly the exception).

The second main requirement for use of any economic instrument to be desirable is that the total combined private and public benefits of the practices being proposed should exceed their total costs.

Economic instruments work by altering the financial incentives and/or risks faced by individuals whose behaviour is targeted (in this case, mainly farmers). The effectiveness of an instrument depends entirely on the strength of incentive it provides and the strength of incentive that farmers would require in order to change practices.

Economic instruments cannot alter the overall desirability of a set of farming practices (from a community-wide perspective). They can only help to increase the adoption of practices that are already socially desirable but are not being adopted, for whatever reason. In effect they redistribute the benefits and costs of biodiversity protection/enhancement such that farmers are given greater incentive to act.

If financial incentives are paid to farmers, they should be less than the monetary value of the resulting non-agricultural benefits (Pannell 2001d). Otherwise the incentives may encourage a level of expenditure by farmers that exceeds the resulting benefits for the community.

Examples of economic policy instruments

Table 3 is a diverse list of types of policy instruments that, broadly speaking, have an economic component. It includes market-based instruments, but also regulatory style approaches, on the assumption that their enforcement is backed with financial penalties. For explanations of the alternative approaches, see ABARE (2001), Pannell (2001d) or Doremus (2003).

Table 3. Examples of economic policy instruments, broadly defined

Source: Pannell (2001d).

Economists have tended to advocate instruments that involve competitive forces (because of the resulting efficiencies), such as market-based instruments or auction-like processes. In practice, however, simpler approaches have been used more commonly. In Australia, by far the most widely used economic instrument for environmental purposes has been partial subsidies for on-ground works (e.g. under the Natural Heritage Trust), conceptually justified on the basis of cost sharing. A more innovative approach has been trialed in Victoria. The approach, in which farmers tender for financial assistance, was briefly described earlier.

Observations on the scope for economic policy instruments for biodiversity

Early evaluations indicate that the tendering process in BushTender can substantially improve the cost-effectiveness of public expenditures (Stoneham et al. 2002). There is enthusiasm to trial this and similar approaches in other places and on a larger scale.

However, the story is not all positive. There is a range of practical difficulties in implementing economic instruments for environmental management, including: the difficulty of choosing a suitable variable to use as the basis for trade or bidding; transaction costs such as the cost of collecting sufficient information to drive the system; choosing the initial allocation of rights; and distributional effects (Pannell 2001d). Furthermore, like any type of policy intervention for biodiversity, economic instruments need to be well targeted; not all proposed biodiversity interventions would pass the positive-net-benefit criterion (or a more subjective version of it that did not require benefits to be expressed in monetary terms).

There is a risk that some policy makers are developing unrealistic expectations about the effectiveness of the more sophisticated economic instruments. I have elsewhere expressed concerns that some senior government agency managers appear to believe that economic instruments are like a perpetual motion machine, achieving the impossible task of getting more out than you put in (Pannell 2001c). Similar concerns appear to have motivated these comments by Boyd and Simpson (1999):

“There is no magic mechanism that will enable conservation at a bargain-basement social cost. [i.e. overall cost to the community, including public and private costs.] What looks like a cheap way of accomplishing conservation objectives is probably a mechanism where costs are hidden, or one where conservation is likely to be ineffective.” (p. 92).

A particular concern is the transaction costs involved in implementing the instruments. These arise from factors such as negotiating with the community about the scheme, administering the scheme, collecting and distributing information, and monitoring and enforcing agreements. For the schemes that are more attractive in theory (e.g. tradable permits), these transaction costs are likely to be high. It is quite possible for the overall transaction costs of potential schemes to exceed its net benefits.

Doremus (2003) proposes that, “Because the options have different strengths and weaknesses, are suited to different contexts, and all carry substantial risk of failure, a portfolio of biodiversity policies is likely to outperform exclusive reliance on any one strategy” (, p. 217).

Economic policy instruments will be most effective in affecting land-use change when the cost of that change is lowest. If it transpires that modest, low-cost changes in management of woody perennials that are planted for commercial reasons can enhance biodiversity, the suitability of using economic instruments to facilitate these changes may deserve investigation.

The Productivity Commission (2001) highlighted the importance of ensuring that existing legislation and regulation does not inappropriately hinder private investment in biodiversity. They identify a number of institutional arrangements that tend to increase the costs and risks of private conservation activities compared with those of other viable land uses. It seems logical to address these impediments with at least the same level of energy devoted to development of new policy instruments.

Private funding for biodiversity

The different economic policy instruments considered above vary widely in terms of who actually pays for the benefits achieved, with different proportional contributions from landholders, the public purse, and private individuals or businesses outside agriculture. Currently, the greatest contribution of finance and other resources for biodiversity conservation in rural areas comes from landholders, followed by public funds, with non-rural private funding making a relatively small contribution.

Environmental organisations are interested in increasing the level of contributions from private non-agricultural businesses, philanthropists and the general public. These groups charge membership fees and regularly make funding appeals for particular causes, and a number of schemes have been attempted or proposed to encourage philanthropic contributions from the business sector.

From a professional perspective, economists view such fees, schemes and appeals with equanimity. Economists generally put a high stake on the sovereignty of individuals and businesses in judging how best to use their own money, and if their judgment is that their money is best spend on biodiversity, so be it. The fact that such schemes do raise funds reflects private preferences for greater levels of environmental protection than is provided by the existing level of public funding.

Economists would caution that we should never rely solely on voluntary private funding for environmental works, because the “non-excludable” nature of some of the benefits (Randall 1981) means that there will be free-riders who benefit without contributing. Some form of government intervention is required to ensure that resource levels are commensurate with the level of benefits generated for the whole community. The form that this generally takes is public funding.


“While society as a whole may benefit from preserving biodiversity, someone must pay to provide those benefits. If habitat is preserved by condemning private land, the private landowner pays. If it is conserved by purchase of private land with public funds, taxpayers pay. If it is preserved via a system of development rights, developers and future property owners pay. Reasonable persons may differ regarding which groups are more morally deserving of bearing or escaping the burden of payment. But someone must pay.” (Boyd and Simpson 1999, p. 92).

This quote begs the question, who should pay? Unfortunately, there is nothing in economic theory that helps us to objectively evaluate the relative merits of different decisions about who pays. For such questions, the contribution of economists is limited to:

A commonly cited rule-of-thumb for distributional questions is the “user-pays principle” or the “beneficiary-pays principle”, under which the beneficiary of a good or service should bear the costs of its provision. It is not a principle in the sense of a scientific principle, but rather a suggestion of what is fair. Ironically, it has been associated with the ideology tagged as “economic rationalism”, although there is nothing in economic theory to favour it as an approach for decisions on distributional issues.

Another commonly cited system for distributing costs is the “polluter-pays principle”. Generally, this approach is in direct conflict with the user-pays approach. It is similarly lacking in any basis in economic theory.

There are problems in trying to rigorously implement either rule. For many environmental issues, it is difficult to accurately identify and quantify the benefits and costs for either the polluters or the beneficiaries of mitigation. The user-pays approach would dictate that members of the community should pay in proportion to their benefits from the biodiversity protection. The polluter-pays approach requires costs to be borne in proportion to the damage caused if biodiversity is not protected. Meeting the information requirements of either rule is impossible in any practical sense, although approximations may be feasible.

Application of any simple rule may be compromised, as governments’ decisions about the distribution of benefits and costs are influenced by a range of considerations. These may include political gain, parochialism, the activities of lobby groups or a wish to benefit particular groups due to perceptions that they are disadvantaged in some way.

The market will also have an influence on the distribution of benefits and costs, irrespective of government wishes. For example, if farmers' production costs go up due to legal requirements to protect biodiversity, the farmers may or may not be able to pass on the increase to consumers of their products. It depends on how responsive consumers are to price changes. If consumers of their products are too responsive, dramatically cutting consumption as prices rise, farmers lose more than they gain by attempting to pass on the extra costs. In a free market, the distribution of costs between farmers and consumers is completely outside government control as it depends entirely on the responsiveness of supply and demand to price changes, and these depend on producers’ cost structures and consumers’ preferences, not on government policy.

In summary, economics offers less help with the question of who should pay than is often suggested. Sometimes there are efficiency dimensions to the question, and economics is certainly useful in addressing these, but usually more important to the community are questions of rights and fairness. These are somewhat flexible over time, driven by community attitudes, politics and power.

An environmental levy?

One aspect that governments can control is the level of public funds allocated to biodiversity. Public funding for environmental and resource management has increased over time, but calls by environmentalists for dramatic further increases are common.

The possibility of introducing an environmental levy, either on the price of food or on income tax, is increasingly being advocated. For example, PMSEIC (2002) identifies as one of its four priorities for investment, “Redressing the absence of economic signals, to urban and rural Australians alike, connecting the underlying ecological condition of natural systems to our use of them for products and services” (p. 14). The currently prominent “Wentworth Group of Concerned Scientists” (2002) notes that in December 2000 a House of Representatives Parliamentary Inquiry into catchment management recommended imposition of an environmental levy, and the Wentworth Group itself suggests a one percent levy on income tax.

However, the general idea is not without problems:

The idea of placing a levy on food is particularly problematic. It could of course be effective in raising revenue, but its efficiency in encouraging better land management is doubtful, and there are several concerns around issues of equity and practicality.

Finally, there are doubts about whether the vastly greater level of funds collected under a levy scheme would be spend well. A critical view of the major national environmental programs in Australia does not inspire confidence that they would.


This paper makes no pretense at presenting a complete discussion of all economics-related aspects of biodiversity. Readers of drafts suggested sufficient additions to expand the paper to a book. Issues that are not discussed include:

Discounting of benefits and costs occurring at different times.


Implications for research

In several parts of the review, the need for improved scientific information was highlighted. There is a critical need to be able to specify the elements of biodiversity that are or should be important to the community, and to be able to justify the selection of these elements. The status and relevance of biodiversity as a policy issue will surely be enhanced if we can clearly express and quantify the benefits of intervention.

Equally important is the ability to predict cause and effect relationships between interventions and desirable outcomes. We need to be able to quantify (or at very least describe) biodiversity benefits of

Given the anticipated increases in commercial plantings of woody perennials in coming decades in landscapes that have traditionally supported grain and livestock production, the third of these areas is perhaps a particular opportunity for beneficial research. This may become a key point of intersection between ecologists and researchers involved in commercial plant development.

Implications for policy

The selection of policy approaches and policy instruments needs to sophisticated, based on science, and sensitive to different biodiversity-related problems and opportunities. For example, this review has identified major differences between the policy responses that are likely to be appropriate for salinity-related and non-salinity related biodiversity issues in Australia.

Where very large scales of land-use change are necessary in order to achieve the desired benefits (e.g. revegetation to control rising saline groundwaters in biodiverse locations), direct payments to landholders or other economic policy instruments appear unlikely to be the most effective or cost-effective approach. The prospects appear better for an indirect approach, based on public investment in development of commercially viable perennials and related infrastructure and industries.

This also generates a potential opportunity to enhance biodiversity values within commercial plantings of woody perennials by beneficial refinements to their management. If not too expensive, these management refinements could be appropriate targets for economic policy instruments.

Such policy tools can also be appropriate for mitigation of non-salinity threats to existing habitat on public or private land. Evidence suggests that careful prioritisation and targeting of the investment to the most cost-effective opportunities is important in order to maximise the resulting biodiversity benefits, and that poorly targeted policy systems can be very wasteful indeed of scarce resources for environmental protection.

The protection of remnant vegetation from clearing was highlighted as likely to be amongst the more cost-effective measures for biodiversity in states where clearing is still permitted. Within this category, vegetation that is not threatened by salinity would appear to offer better prospects for cost-effective public investment. Perhaps a caveat on the conclusion in favour of reduced clearing could relate to the transaction costs arising from possible political resistance if changes are made compulsorily.


Thanks to Ted Lefroy for the invitation to prepare this paper, and to Ted, David Freudenberger and Kim Lowe for providing stimulating ideas and advice on its content. I appreciate feedback on early drafts provided by Deborah Peterson, Kim Lowe, Jason Alexandra, Ben White, Neil Byron, Lesley Brooker, Michael Brooker, Geoff Edwards and Daniel Spring.

The CRC for Plant-Based Management of Dryland Salinity and Meat and Livestock Australia supported the workshop and the Grains Research and Development Corporation funded past research that helped inform this review.


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This paper was prepared for and presented at a workshop of the Cooperative Research Centre for Plant-Based Management of Dryland Salinity, “Biodiversity Values in Agricultural Landscapes”, Rutherglen, Victoria, 14-15 October 2003

Citation: Pannell, D.J. (2003). Heathens in the chapel? Economics and the conservation of native biodiversity, Presented at a workshop of the Cooperative Research Centre for Plant-Based Management of Dryland Salinity, “Biodiversity Values in Agricultural Landscapes”, Rutherglen, Victoria, 14-15 October 2003. http://www.general.uwa.edu.au/u/dpannell/dp0301.htm

A modified version was later published as:

Pannell, D.J. (2004). Heathens in the chapel? Application of economics to biodiversity, Pacific Conservation Biology 10(2/3): 88-105.

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