When environmental pollution became a pressing concern in the industrialized countries in the late 1960s, the economics profession responded with analyses pointing to flawed incentives based on an incomplete specification of property rights and the concomitant failure to guide socially efficient behavior through market processes. (See, for example, Ayres & Kneese, 1969; Dales, 1968; and Ruff, 1970.) Economists aspired to correct the incentive structure by establishing markets for pollution. Although early proposals for effluent charges and tradable emission permits contained important insights, they were flawed because they were based on oversimplified models of both the physical and social dimensions of environmental pollution and regulation. When concrete policies were being established, economists' proposals largely were ignored in favor of what has come to be called the command-and-control (CAC) approach, i.e., direct regulation based on rigid specification of pollution control requirements. By the early 1970s virtually all the industrialized countries had some form of environmental authority, and while some of the early policies had economic incentive features, they all relied foremost on direct regulation.
Research through the intervening decades has substantially improved understanding of both the physical properties of pollution and social responses to it. This progress is reflected in proposals for differentiated market instruments that account for the importance of location of emissions, work on environmental valuation in support of cost-benefit analysis, analysis of the redistributive impacts of environmental policy, and explicit consideration of implementation problems of political opposition and administration (Cropper & Oates, 1992). In particular, economists have recognized that ignoring the complexity inherent to pollution often results in proposing inefficient or unworkable policies. At the same time policymakers and environmental activists have recognized desirable aspects of economists' proposals, including their flexibility, incentives, and efficient use of local information, that can improve the cost-effectiveness of environmental policy and stimulate technical innovation in pollution control.
Although never completely ignored, applications of market instruments have increased as environmental regulatory systems have evolved, and recently them has been enhanced interest in this approach to pollution control (Keohane, Revesz, & Stayins, 1997; Oates, 1994; Oosterhuis & de Savornin Lohman, 1994; Opgchoor, de Savomin Lohman, & Vos, 1994). Increased awareness of ecological problems in developing countries has led them toward environmental policies consistent with economic development plans, and sustainable development initiatives in a variety of countries explicitly have recognized market-type instruments as useful tools in environmental policy (Turner & Opschoor, 1994). Moreover, the economic transition of formerly Communist countries and revelations of their extensive environmental degradation raise questions of how they should restructure their environmental policy. These are all reasons to review the use of market-type instruments for environmental goals and consider lessons from this experience.
This paper does not present a comprehensive survey.(1) The countries considered are limited to five industrialized capitalist nations (France, Germany, Japan, the Netherlands, and the United States). Each has a well-developed system of environmental regulation, and the free-market orientation of these societies is most likely to be receptive to market approaches for environmental regulation. In the interest of brevity, the scope is limited to problems of common air and water pollutants. The market instruments considered include emission charges, subsidies for abatement, and tradable emission permits. The common elements among these include a devolution of decision authority to firms emitting pollutants and monetary payments based on their behavior. A recent analysis of the political economy of environmental policy (Keohane et al., 1997) notes differences between the United States and European countries in their application of market instruments. The latter have emphasized emission charges while the United States has moved toward tradable permits. The comparative approach of the present paper leads to some specific conclusions about this observation that are tied to differences in the culture of enterprise regulation.
The next section reviews the conceptual framework of market instruments and is followed by an examination of policies in the five countries in section 3. Section 4 attempts to reconcile theory with practice by expanding the analysis of market approaches beyond the narrow economic considerations discussed in section 2 to include other dimensions of environmental policy. The paper considers how these features have affected adoption of concrete policies for environmental control. A final section presents conclusions.
Theory of Market Instruments
The theory of market instruments for pollution control has a substantial history. One of the original sources for this line of thinking is Dales (1968), who proposed a system of tradable permits, and the taxation idea for correcting externalities can be traced back to Pigou (1938). An elaborate literature exists on taxation for externalities, and one of the best summary statements of the theory in a general equilibrium context is Baumol & Oates (1988). The fundamentals of the theory of market instruments can be found in any textbook on environmental economics. (See, for example, Callan & Thomas, 1996; or Lesser, Dodds, & Zerbe, 1997.) To contemplate an optimal level of polluting emissions, an environmental authority must have knowledge of marginal environmental damages, and this is difficult to acquire. Indeed this problem has become a field of study in itself. (See, for examples, Braden & Kolstad, 1991; Freeman, 1993; and Johansson, 1987.) Because of this, actual environmental policies have not balanced cost and benefit at the margin, as optimality requires.
However, cost-effectiveness for a selected environmental target is more tractable. In simple models, market approaches lead to least-cost abatement solutions. Firms select cost-minimizing control technology, and the marginal cost of abatement is equalized across firms, implying efficient allocation of abatement activity. Moreover, incentives to spur technical change in pollution control are inherent. The lack of these features in CAC policies is the main criticism by proponents of market approaches. The essence of the market approach to pollution control is to force firms to bear a cost tied to their emissions while allowing managers to make independent decisions regarding control. The connection of cost to emissions is seen most directly in the emission charge policy, but it is present under other market instruments as well. A CAC approach usurps this independence and sets specific guidelines for emitting pollutants. Thus the market approach alters the market circumstances that guide firms' decisions, while the CAC approach restricts the scope of firms' decisionmaking power.
Command versus Market
While often juxtaposed as opposites in policy discussions, the market and CAC approaches are fundamentally similar in restricting a firm's ability to exercise property rights to the waste disposal capacity of environmental media. Under a CAC policy such rights are assumed to lie with polluting firms but are subject to extensive attenuation by the environmental authority. Under the market approach there is little or no attenuation, but the right must be acquired explicitly, typically via a market transaction.(2) The distinction between market and CAC approaches is blurred further by the ability of the environmental authority to attenuate not only a recognized right but legitimate acquisition itself. In some U.S. programs a firm's use of purchased emission rights for compliance is allowed only if the firm is in compliance with technological standards.
Despite some ambiguity, the distinction facilitates policy analysis, and the state, through its environmental authority, has distinct but central roles in both policy approaches. Under CAC this role is to establish and enforce appropriate attenuations. Under a market approach it is to establish a procedure and legitimacy for rights acquisition, just as it does in other areas of exchange in a market economy. The state's basic function here is to resolve social conflict arising from the use of environmental media. Markets are one means of mediating conflict, but other institutions can serve this purpose, sometimes more effectively. The wide potential for degrees of property rights attenuation, for example, confers a great deal of flexibility to CAC regulation that frequently is overlooked by proponents of market approaches.
The basic similarities between market and CAC approaches suggest the same outcome could be achieved under either. While in principle this is true, the associated information processing tasks are very different. From this perspective, the main advantage of market instruments is that they simplify the information problem faced by an environmental authority without complicating it excessively for regulated firms. The two marks of cost-effective outcomes are selection of least-cost technology and allocation of abatement responsibility in a way that equalizes marginal abatement cost. A cost-effective CAC policy requires the environmental authority to acquire knowledge of least-cost technologies as well as the marginal cost conditions of all firms that will be assigned an allowable emission level. Market-based policies process the same information but in a decentralized manner that may accomplish the task more efficiently. This amounts to a specific application of the general principles of information processing in complex systems (Hayek, 1945; Stiglitz, 1994).
The typical comparison between "rigid" CAC approaches and "flexible" market approaches arises because an environmental authority, overwhelmed by the information requirements, establishes uniform regulations that, when applied to nonuniform circumstances, lead to excessive cost in meeting environmental targets. Certainly this characterizes many environmental regulatory regimes. However, where the information processing problem is tractable for a centralized system, arguments against CAC approaches are weakened substantially. This is important in comparing environmental policies across countries. Regulated firms possess much of the critical information needed, and where cooperative relations have evolved between firms and regulators the information problem is diminished and enforcement problems are mitigated. Cultural expectations about the role of the state in economic affairs influences the nature of these relations.
Physical features of ecosystems preclude achieving least-cost solutions using uniform market instruments. Because location of emission matters for ambient conditions, and ambient conditions are the policy objective, the location of emission is important for regulatory policy. For charge and tradable permit systems to be cost-effective they must be differentiated spatially (Kolstad, 1987; Montgomery, 1972; Tietenberg, 1978). When multiple receptor sites are considered, multiple systems of transfer coefficients are required, and the goal of cost-effectiveness is reached only by a very complex charge or tradable permit system. The establishment of a regulatory regime that would take advantage of the market's ability to process information efficiently would become itself a highly information-intensive undertaking. Several other considerations complicate the theory of market instruments, reducing the attractiveness of their erstwhile simplicity. One is uncertainty in abatement costs and environmental benefits.(3) Performance of market instruments under conditions of change (e.g., economic growth or inflation) is another. The significant income distribution shifts that market instruments can entail reduce their appeal, as coalitions form in opposition to them. Moreover, the ethical foundation, that polluting is acceptable if dues have been paid, may diminish public support for these regulatory strategies. Finally, a major reason for reluctance to adopt effluent charges is that indirect control over pollution is perceived as unlikely to attain the environmental quality goals society has established (Bohm & Russell, 1985).
Variants of Market Instruments
Per-unit subsidies theoretically are similar to effluent charges. Here society grants emission rights to firms and then buys them back at a set price if they are not exercised. Although there are long-run efficiency reasons for avoiding per-unit subsidies, the main reason in practice is the equity dimension and adherence to the principle that the polluter pays.(4) This principle requires those who generate pollution to bear the cost of controlling it. Although firms themselves would make abatement expenditures, an offsetting subsidy shifts the cost burden to the general tax base. Such subsidies have not been applied as pollution control instruments.
Other forms of subsidies, however, are used widely to support capital expenditures on pollution control equipment. Across Organization for Economic Cooperation and Development (OECD) countries, around 5% to 20% of all investment in pollution control is paid for by subsidies (Opschoor & Vos, 1989). These subsidies are not market instruments; they neither create nor mimic markets in emission rights. Consequently, they lack the desirable incentive features of market instruments. Yet subsidies are very important in addressing the distributional burden of environmental policy and in engendering policy support among the actors who will be regulated. Recognizing their wide application, the OECD countries have agreed to allow them where severe financial difficulties would arise otherwise, for transitional periods, and in cases where international trade and investment are not significantly distorted (Opschoor & Vos, 1989).
Both product charges and quasi-tradable permit systems have been applied to pollution control. Although different in some respects, these instruments are much like effluent charges and tradable permit systems, respectively. A product charge is a tax on a product that generates pollution at some stage in its production-consumption cycle. The higher price induces users to substitute other products, thus reducing emissions of the related pollutant. This tool is useful where it is costly to monitor actual emissions, as in automotive fuels. Product charges intended as an incentive can be established also as a tax differential between two substitute commodities, with a lower tax applied to the one causing less environmental damage. The United States Environmental Protection Agency's (EPA's) offset and bubble programs are examples of quasi-tradable permit systems. Unlike a pure permit system, explicit property rights are not issued, but rather arise when a firm demonstrates that it has reduced emissions below an assigned standard. Subject to verification by the environmental authority, these "emission reduction credits" become transferable commodities analogous to a tradable permit in the pure system.
Practices in Five OECD Countries
From this theoretical perspective, we turn now to consideration of the practical experience with market-type instruments in five industrialized countries. Per capita incomes and efforts at pollution control are roughly similar across these countries. In 1990, gross domestic product (GDP) per capita ranged from $19,000 in the Netherlands to $25,000 in Japan (International Monetary Fund, 1995). In 1995, this ranged from $20,000 in the Netherlands to $26,000 in the United States (Organization for Economic Cooperation and Development, 1997).(5) Each country has a significant regulatory structure devoted to environmental protection based on major legislation established over the past three decades. Estimates of annual expenditures on pollution control in 1989 and 1990 ranged from 1% of GDP in France(6) to 1.6% of GDP in Germany (Kopp, Portney, & DeWitt, 1990; Organization for Economic Cooperation and Development, 1993).(7) Although there are differences across countries, none is especially striking. In some, significant funds are raised from effluent charges, but only in the case of the Netherlands do these comprise more than a few percent of total expenditures on pollution control.
Important characteristics of the French regulatory regime include a high degree of centralization and extensive cooperation between regulators and firms. Economic development and environmental regulation are seen as closely linked, and both are within the province of state authority in keeping with the French tradition of indicative economic planning. The French experience with market instruments is limited to two effluent charges and the provision of subsidies through grants, soft loans, and accelerated depreciation allowances. Much of the subsidy support is linked directly to the charge revenues, and both charge systems operate in conjunction with CAC regulation.
The water effluent charge (begun in 1964) is administered by six basin authorities, the Agences Financieres de Bassin, that determine the level of charges and plans for expenditures on water quality projects financed by the revenues. These policies are formulated jointly by the basin authority and industrial enterprises in its region (Feldman, 1989). By law the agencies are revenue neutral, meaning that the charge in each watershed is determined by planned expenditures. All actors emitting polluting effluent are liable for the charge, which is based on the volume and physical characteristics of the effluent. Revenues collected go to three classes of expenditures: administration of the agencies, public expenditures for water quality projects, and subsidies granted to enterprises to support investment related to pollution control. Although the effluent charges have some incentive effect, the main purpose is to raise revenues. The levels are very low compared to what an incentive charge would have to be to reach the environmental target specified in the CAC regulations (Opschoor & Vos, 1989).
The air pollution charge, instituted in 1985 and renewed in 1990, follows the same basic structure but is administered by the centralized Air Quality Agency. Pollutant coverage includes all acidifying emissions from power plants and large industrial firms. Around 90% of the revenue is returned to firms to finance abatement investments; the remainder supports the Air Quality Agency and its research and development (R&D) programs. The air pollution charge also is too low to have enough incentive impact to reach environmental goals. Without direct regulation the charge would have to be around two orders of magnitude larger to bring the firms into compliance with European Community guidelines (Opschoor & Vos, 1989). Although the charge increased in the 1990s, it has not risen to such a level.
The main features of these programs are their self-financing nature and their combination with direct regulation. The charges are a way of generating revenues for environmental programs and of maintaining consistency with the "polluter pays" principle while extending subsidies to industrial firms. Industrial interests oppose attempts to increase the charges, and they would find charges at the incentive level unacceptable. Environmental goals are pursued using direct regulation, while supporting industry with subsidy programs tied to investments in pollution control that were important in reducing industry opposition to the emission charge systems (Opschoor & Vos, 1989).
Similar to the French case, the formulation of German environmental policy involves considerable negotiation with regulated firms, but German regulation is more decentralized, with the Lander (state governments) and municipalities playing important roles in policy formation (Brown & Johnson, 1984). Moreover, the Lander and local governments have primary responsibility for implementation. Conflict between regulators and firms is mitigated by a common philosophical outlook on the government's role in the economy (Miller, 1989). Intervention in an otherwise free market economy for the purposes of promoting the general social welfare is a position accepted by all the major political parties in Germany.
The wastewater effluent charge, instituted in 1981, was developed through a great deal of negotiation between federal authorities and firms, and between federal authorities and the Lander. It is tied closely to a system of effluent standards and permits. To emit effluent, a firm or municipality must have a permit that designates an expected volume and concentration. Charges, differentiated by type and concentration of pollutant, are applied to all effluents, which are measured in a complex system of monitoring. The charge assessed depends on the extent of a firm's compliance with the regulatory standard, and this is where the central incentive feature of the German system comes to bear. If a firm reduces emissions below its assigned standard, it qualifies for a lower rate. If a firm consistently exceeds its emission standard, its permit will be renegotiated to reflect a higher volume of effluent flow. Although the incentive structure of the charging system has led some firms to alter behavior so as to qualify for the charge reduction (Opschoor & Vos, 1989), attaining environmental quality goals relies foremost on direct regulation (Brown & Johnson, 1984). Charge revenues go to the Lander to cover administration, public expenditures on related R&D, and pollution control, and subsidies to private industry for investment in control equipment. Only a small part of these subsidies is funded by the revenues collected in the wastewater charge system.
The application of emission charge policies for air pollutants in Germany has been limited to tax differentials on leaded versus unleaded fuels and on high versus low-emission automobiles. Germany also has established programs of limited trading in emissions similar to the U.S. offset and netting provisions. In areas not reaching the ambient air quality standard, new sources of emissions are allowed only if a firm can obtain reductions in emissions elsewhere in the area. Reductions can come from within a firm or from interfirm negotiations. The compensation settlements program is like the U.S. netting provision. Because compensations must occur within restricted geographic areas, this policy is of limited importance (Opschoor & Vos, 1989). German subsidies support industry expenditures on pollution control equipment. These programs are meant to provide financial support to firms that would be stressed unduly by the requirements of environmental regulation and to accelerate implementation of environmental quality programs.
Although German policies have market elements, basically they rely on a system of direct regulation within which the market instruments act as supplementary features to lend a degree of flexibility in firm adjustment to environmental standards (Opschoor & Vos, 1989). This flexibility appears to have improved cost-effectiveness significantly. The Council of Experts for Environmental Questions estimated that the hybrid system for water quality reached the environmental goals for about two-thirds the cost that would have been incurred under an inflexible uniform standard (Brown & Johnson, 1984).(8)
As in the case of France, the government in Japan plays a significant role in economic planning, and this tradition has a bearing on how environmental policy is approached. In the postwar era the main objective of economic policy has been to foster growth, often at the expense of environmental quality. Japan's high population density and advanced degree of industrialization created some very serious pollution problems (Marcus, 1989). With the rise of public complaint over the state of the environment (Marcus, 1989; Miyamoto, 1991; Totsuka, 1989) and the appearance of acute illnesses associated with pollution, an environmental control program was forced on the pro-growth planners. Yet the overriding philosophy was that environmental goals should be harmonized with economic growth and not pursued at the latter's expense. There is extensive interaction with firms in formulating and implementing environmental policy, and these relations are characterized as cooperative. "Emission standards are enforced by persuasion rather than coercion" (Totsuka, 1989, p. 331). While policy frameworks are decided by the central government, implementation is conducted mostly by local governments negotiating with polluting firms in their area to specify how the regulatory standards are to be applied in each instance. Local authorities have significant decision power regarding pollution control (Totsuka, 1989).
Japanese evolution of environmental policy proceeded somewhat differently from that of Europe, having an emphasis on the compensation of pollution victims. A galvanizing event in the Japanese experience was the linkage of some very serious diseases with heavy metal poisoning.(9) A series of law cases in which victims sought and won compensation were tried in the late 1960s and early 1970s (Miyamoto, 1991). To avoid continuing litigation costs, in 1973 the government passed the Pollution-Related Health Damage Compensation Law. Persons who could show that they suffered from diseases known to be linked with environmental pollution automatically were granted financial support from the relevant compensation fund (Totsuka, 1989).
The fund for victims of air pollution was established on the basis of an emission charge(10) applied to major industrial sources of S[O.sub.2]. These sources accounted for around 90% of total emissions (Baumol & Oates, 1979). The size of the charge was calculated to meet the financial obligations of the fund; thus it had no relation to abatement costs. This charge was in place from the mid-1970s to 1988, when it was discontinued under pressure from business interests. When the program was ended there were over 103,000 persons receiving 100 billion yen in compensation (Miyamoto, 1991, p. 674). A number of indirect subsidy programs have been utilized in Japan. One lowers the commodity tax for low-emission automobiles. Another reduces the customs duty on low-sulfur crude oil. Industrial firms receive tax concessions for investments in pollution control equipment, including accelerated depreciation rules and tax credit for up to one-third of expenditures on control equipment (Marcus, 1989).
The Netherlands has one of the most progressive and comprehensive environmental policies in the world and significant experience with market instruments. The approach to administration and formulation of policy is consultative and inclusive. Dutch officials seek to engage economic sectors in negotiations, leading to covenants on environmental objectives. Subsequently, these covenants are enacted into law and form the legislative basis of environmental policy (World Resources Institute, 1994). Environmental programs have followed the Dutch philosophy of establishing taxes that, as closely as possible, place the fiscal burden of government programs on the economic actors that create the need for public expenditures (Oosterhuis & de Savornin Lohman, 1994). The Netherlands is similar to Japan in having a high pollution load per unit of land area, owing to intensive industrialization and high population density. In the early decades following World War II it was considered one of the most polluted countries in the world. There is significant damage from an agriculture with limited arable land and intensive use of chemicals. The Dutch manure problem causes water pollution, and in the late 1980s policies specific to this were adopted, including application of market instruments.
The Netherlands has been successful in controlling water pollution through a combined charge-standards system established in 1970. Firms pay charges on effluents according to volume and composition to local water boards (waterschappen), who in turn spend the revenues on water quality programs including R&D, public treatment facilities, and subsidy programs for private enterprise. A high degree of compliance is attributable to the waterschap, a uniquely Dutch institution having roots in the Middle Ages (Huppes & Kagan, 1989). These agencies have administered water policy and levied taxes for generations, and they enjoy strong legitimacy among the public. The system is revenue-neutral, so the size of the charge is related to the expenditure of the waterschap. Dutch policy is characterized by decentralized administration and a high charge level, imposing a significant incentive impact on large firms. Small firms and households pay flat rates unrelated to their levels of emissions. For perspective, annual per capita revenues from water effluent charges (in the mid-1980s) were $38.9 in the Netherlands as against $4.6 and $2.3 in Germany and France, respectively (Opschoor & Vos, 1989). During the policy discussion stage the charges were put forward not as incentives, but as an equitable way to finance the sorely needed investments in water pollution control (Huppes & Kagan, 1989). The incentive impact subsequently was studied by Bressers (1988, p. 509), who calculated a correlation coefficient between charges and pollution abatement of .86.(11)
The manure policy (1987) has a number of components. Application of manure to fields in excess of a stipulated upper level is subject to a progressive waste tax. Two related product taxes have been levied: one on imported fodder and the other on farm animals exceeding a designated number of head per hectare. The impact of this program is significant for animal-intensive operations such as modern hog and poultry producers. While the industrial charge-standard system proved very successful, the start of the manure program was less auspicious (Huppes & Kagan, 1989). Farmer opposition was significant, the manure accounting systems were complex, and monitoring was a problem. The explanation for the difference in success may lie in the institutions of implementation. The waterschappen are local authorities dominated by landowners, while the manure policy is administered by the central government's Department of Agriculture (Huppes & Kagan, 1989).
Some other market instruments in the Netherlands bear mentioning. In 1988 a general fuel charge replaced five separate product charges to provide financing for the Ministry of the Environment (Opschoor & Vos, 1989), and the Dutch are among the few countries in the world to have levied a carbon tax on fossil fuels. There are two subsidy programs. One assists firms that otherwise would be stressed severely by environmental compliance. The other supports R&D in pollution control and clean production processes. There has been considerable discussion in the Netherlands on tradable permit systems, but none has been introduced yet. Limited provisions in air quality regulations allow trading in emission reductions similar to the bubble policy in the United States, and reduction offsets often are considered in negotiations between regulators and firms, but there is no formal program for trading in emission rights. A ban on leaded automotive fuel was implemented gradually, and to facilitate a more rapid transition a tax differential between leaded and unleaded fuel was established as an incentive to consumers. This appears to have been redundant because the transition to unleaded fuel occurred more rapidly than expected (Oosterhuis & de Savornin Lobman, 1994).
The administrative system for environmental policy in the United States is a mix of centralized and decentralized. While states have a great deal of independence and primary responsibility for implementation and enforcement, environmental standards typically are determined at the federal level by the EPA. Among the countries studied, the United States has the most confrontational regulatory tradition, reflected in the extent of litigation brought by firms in challenging the regulatory authority (Webber, 1989). The philosophy of regulators has been tied closely to the CAC framework, including specifications of particular pollution abatement technology, an extent of control generally not practiced by the European countries. A more intrusive approach to regulation of private enterprise in the United States has been noted by several researchers of comparative politics, and in particular by Vogel (1986) for the case of environmental regulation.
The use of emissions charges has been very limited in the United States, being restricted to wastewater effluents. These charges are administered locally, their primary purpose is to fund treatment plants, and usually they are not tied directly to effluent volumes. As incentive impacts are negligible, these charges don't really represent application of market instruments for pollution control. In keeping with a philosophy contrary to government support of private industry, the only significant subsidy program in U.S. environmental policy has been intergovernmental transfers in the form of federal grants to local governments for the construction of wastewater treatment facilities. The evolution of market instruments in the United States has been limited mostly to tradable permit systems for air pollutants, although the Clinton administration has moved to widen their scope in environmental policy (Cushman, 1995).
Two tradable permit systems pertaining to water pollution were established under local authorities: the Fox River and the Dillon Reservoir programs in Wisconsin and Colorado, respectively. Both programs supplemented CAC policies. In 1981, the Wisconsin Department of Natural Resources issued permits and established rules for trading in biochemical oxygen demand, but there has been essentially no trading. Opschoor & Vos (1989) speculated that the main cause was firm competition in product markets (a series of pulp mills are the main object of regulation on the Fox River), but others argued more convincingly that stringent administrative requirements were to blame. These, in turn, resulted from the highly charged political environment that characterized this policy evolution (Hahn & Hester, 1989). The Dillon Reservoir program established trading provisions in 1984 for phosphorus in an attempt to stimulate greater control from nonpoint sources. Lack of trading activity and subsequent mitigation of the reservoir's eutrophication problem have rendered this program irrelevant (Hockenstein, Stavins, & Whitehead, 1997).
In the late 1970s the EPA began to experiment with flexibility in emission standards for air pollutants, and legislation in 1986 formalized procedures for quasi-tradable permit systems referred to as the bubble, offset, netting, and banking provisions. Banking allows a firm temporarily to store emission reduction credits, and the other three elements are different applications of tradable permits. The offset program allows trades between existing and new sources, while the bubble policy provides for trades among existing sources. The netting program pertains to equipment modifications within a given production facility. Several observers argue that these programs are not as successful as they could be. Although their introduction has moved policy toward tradable permit systems, they have been criticized as unduly restrictive and excessively bureaucratic. Trades are subject to extensive regulatory reviews, approval typically takes from 4 to 29 months, and trade brokers have emerged in response to the high cost of trading (Opschoor & Vos, 1989). Of the several thousand offset trades completed by the late 1980s only 100 were at "arm's length" (Dudek & Palmisano, 1988), which suggests that administrative requirements are biased in favor of internal shifts of abatement responsibility. High transaction costs likely are preventing mutually beneficial trades between separate firms that would reduce overall compliance cost.
Despite limitations on trading, these policies have helped achieve major cost savings. With the flexibility they introduce, firms are less tied to particular control technologies, which is especially important because policy in the United States has been technologically restrictive. There also is scope for shifting abatement burdens from high-cost to low-cost sources. The extent to which limited trading is able to exploit the potential cost saving is not known fully, but several studies have indicated that significant reductions in compliance cost have resulted (Dudek & Palmisano, 1988; Hahn & Hester, 1989; Liroff, 1986). Hahn & I-lester (1989) estimated over $400 million of savings by the 132 trades that had occurred under the bubble program and as much as several billions of dollars by the thousands of trades carried out under the netting program. Their analysis concluded that the following features are important for success: low levels of political conflict, low information requirements for firms, lessened regulatory delay, security in the property right, and the existence of facilitating institutions that function as a trading infrastructure. These features are basic to economic exchange more generally.
Two minor programs of note are the lead and chlorofluorocarbon (CFC) trading provisions, both of which were utilized extensively by firms in the respective industries. Both policies were designed to provide a degree of flexibility for firms facing severe restrictions on the production of leaded gasoline and CFCs, respectively. As such, the trading programs were complementary to strict CAC policies and have ended with the passing of the phaseout dates. The lead-trading program was a quasi-tradable permit system much like the bubble provision. Research indicates it was successful in reducing the cost of transition (Hahn & Hester, 1989). The CFC program was closer to a true tradable permit system, but had significant restrictions imposed. Each transaction required EPA approval, and a producer could obtain increases only up to a maximum of 15% of its allowance (Tietenberg, 1996).
A new stage of tradable permit policies arrived with the implementation in 1995 of the Allowance Trading Program for S[O.sub.2] emissions. Allowances were allocated according to historic emission patterns except for a small portion (2.8%) held back by the EPA for annual auction by the Chicago Board of Trade (L. Montgomery, October 15, 18, 1994, personal communication; Mr. Montgomery is an EPA administrator in the Acid Rain Division, Washington, DC). The second phase of the program will broaden coverage to all electric utilities in the United States. While permit trading is concentrated within the power industry, firms from other industries with S[O.sub.2] restrictions can participate. Industry response to the program is providing an interesting laboratory for economists. Although trade volume and permit prices have been lower than predicted, the market is active, and estimates of the annual saving in compliance cost are around $1 billion (Cushman, 1991; Hockenstein et al., 1997).
An intriguing recent development in U.S. policy is the establishment of emissions trading programs on the initiative of state and local governments. California's South Coast Air Quality Management District has been notably progressive with its program for the Los Angeles basin (implemented in 1994), but there also are new tradable permit programs in early stages of implementation in Michigan and the tri-state area of New York, New Jersey, and Connecticut. The Los Angeles market is quite active (Hockenstein et al., 1997), suggesting better prospects for success of emission trading in air pollutants than foretold by the limited experience with wastewater programs at the local level. If this prognosis is borne out, tradable permit policies administered at the state and local level are likely to become a policy trend.
In summary, water pollution policy in the United States has remained essentially one of CAC. Despite the Fox River and Dillon Reservoir experiments, practically speaking, tradable permit systems for water pollutants have not been developed. Charges for wastewater emissions are limited to raising funds for treatment programs. Tradable permit programs in air pollutants, on the other hand, are significant. Despite restrictions, a great many trades have occurred, and significant savings in compliance costs have been realized. With the advent of the S[O.sub.2] allowance trading and new programs administered at the state and local level, a new stage in the application of market instruments has been reached.
Practice versus Theory
Although governments have been engaged in widespread environmental regulation for several decades, reliance on markets for pollution control is quite limited. Application of market-type instruments generally is different from the economist's vision. Subsidies are common, but limited to lump-sum transfers supporting investment in control equipment. Emission charge systems are instituted primarily to raise revenue for environmental programs rather than as incentive policies (Opschoor et al., 1994), with charge levels based on projections of financial need to fund designated programs. Any charge presents an economic incentive to avoid emissions, but none is high enough to be an effective control by itself. Environmental goals are attained with the use of direct regulation. Tradable permit systems, which are limited to the United States, often have imposed extensive constraints on trading, reducing their capacity to function as market instruments. Programs of tradable permits in water pollutants have not had any impact. The S[O.sub.2] allowance trading policy and emission trading in the Los Angeles airshed are the most progressive toward true market-based regulation.
In social systems open to policy experimentation it is reasonable to expect an evolution toward regulatory approaches that are more efficient. Only the United States exhibits a significant tendency toward market-based pollution control, and development there has been notably slow. This observation suggests the lack of a compelling advantage inherent to market instruments. Explanation of this situation seems to lie in two areas: the political nature of the regulatory problem and the flexibility of CAC approaches.
Politics of Environmental Regulation
Environmental regulation is formulated in a political process, and public attitudes depend on how private wealth will be affected and on citizens' ethical points of view on the environment. The fundamental social problem of pollution control policy is resolving the associated conflict, and market approaches are not always the most effective mechanism of conflict resolution. Since the smooth functioning of markets more generally is facilitated by social consensus, market approaches to environmental regulation will prove more effective where the basic conflicts over environmental resource allocation have been resolved. Under certain conditions, first outlined by Coase (1960), the spontaneous development of markets may accomplish this conflict resolution. But for many environmental problems the conditions required by the Coase theorem are not in place, and only administratively developed markets can arise. Direct regulation rather than synthetic markets may be preferred because policy formulation of this type is less conflictive. Details of direct regulation typically are established in an administrative arena, whereas details on charge levels and scope of application must be developed in a political arena that is inherently more contentious (de Savornin Lohman, 1994). Realistic environmental policy, then, must be compatible with political circumstances. Turner & Opschoor (1994) have formulated this observation as a "concordance principle" that requires a regulatory approach to be in accord with the political and administrative situation in which it will be applied. Ethics, distributional consequences, and administrative capacity all play roles in determining this compatibility.
Two broad ethical viewpoints on environmental policy can be distinguished: instrumentalists, who see the environment as a tool for improving human welfare, and environmentalists, such as the "deep ecology" movement, who see preservation of the environment as an end in itself. An effluent charge, representing a price payable for the right to pollute the environment, is acceptable in principle to instrumentalists, although they are concerned with effectiveness in reducing emissions and distributional implications, but the idea may offend an environmentalist. In their belief system there is no price that can be placed on the environment. Both ethical viewpoints are expressed in the policy process, and measures adopted represent a compromise mediated by the political system. In view of the ecological practices in the countries surveyed, it is apparent that the instrumentalist position is dominant, in that polluting emissions have been allowed, but environmentalists have resisted market-type controls that convey legitimacy to polluting. Thus ethical viewpoints and suspicion of the environmental effectiveness of market-type instruments have worked against their adoption (Bohm & Russell, 1985). Proof of this for the United States can be found in Kelman's (1981) surveys of congressional staff on their views of the application of emission charges. The active participation of environmental groups in policy formulation and their negative assessment of market-type instruments is further evidence of the practical impact of ethical stances in policy formulation (Keohane et al., 1997).
The distributional implications may be more important than ethics, however. Although the long-run incidence of the cost of pollution programs (including pure transfers) likely will be quite different from initial impacts, the latter most often are the focus of political debate (Organization for Economic Cooperation and Development, 1994a). Private industry (which surely has little environmentalist ethic) has resisted regulation by emission charges because they would impose very large transfers of wealth to the environmental authority. In tradable permit systems such a transfer would occur if permits were allocated initially by auction. Where charge systems have been imposed, the charge levels result in much smaller transfers than would occur under incentive-based systems. Moreover, acceptance of charges at even these lower levels has been facilitated by subsidies for pollution control. In tradable permit systems there never has been an initial allocation by auction. Permits have been awarded to firms according to historical patterns of emissions.
In an early effort toward developing a theory of policy choice, Buchanan and Tullock (1975) stressed the distributional consequences of CAC in comparison to emission charges, but they also noted a possible reluctance by policymakers to usurp property rights by legislation, which the imposition of charges effectively does. Although legislative attenuation of rights is inherent to CAC regulation, it is not as extreme. Firms would perceive the imposition of charges as confiscatory, and so oppose them in favor of direct regulation. Opposition to charges could be mollified by a policy of extending lump-sum transfers to firms as compensation for capital losses they would incur. This analysis is consistent with the observed development of environmental policy. Although no country has imposed charges at a level required for pollution control without complementary CAC regulation, countries that use charges also provide subsidies to industry.
The force of the distributional issue also is evident in the practice of earmarking charge revenues for environmental expenditures. Although it is appealing politically to use them to fund environmental programs, that is not necessarily their best use. Fiscal efficiency might be improved by using charge revenues in nonenvironmental programs. This brings environmental policy to the larger question of fiscal hypothecation, which cannot be treated here, but it is important to note that emission charges at levels required to have sufficient incentive effects to meet environmental targets would generate revenues well in excess of public spending on pollution control. This approach could be made acceptable politically only if taxation in other areas (e.g., income and value-added taxes) are reduced accordingly, which would require a thorough revamping of fiscal structures unrelated to environmental policy. Proposals have been developed for environmentally oriented reforms of entire taxation systems, and Sweden has begun to implement tax changes in this direction (Organization for Economic Cooperation and Development, 1994b).
Cost-Effectiveness of Command-and-Control
One of the main criticisms of CAC policies is their lack of flexibility, yet implementation of a CAC policy can incorporate a great deal of flexibility. A policy based on emission limits (Germany and the Netherlands) rather than on technology specification (the United States) allows firms to choose control technology. Negotiation to alter emission targets allows a CAC policy to be adjusted to local conditions, which is a feature of the Japanese system. Correctly specified, direct regulation can approach cost-effective results. The superior dynamic efficiency of market approaches also can be questioned. Although a firm directed to apply a particular technology has no incentive to innovate, one that is directed to reduce emissions will seek out least-cost abatement technology as an aspect of overall cost minimization. Furthermore, the private sector is not the sole source of research and development in pollution control. As in all areas of industrial technology, public sponsorship of research has made important contributions to innovation in pollution control. Moreover, the impact of economic incentives on abatement innovation has not been clearly established (Turner & Opschoor, 1994).
The culture of public administration plays a pivotal role in the application of CAC policies. Regulatory culture, by impinging on the costs of enforcement and information processing, influences the extent to which CAC policies can achieve cost-effective results. In capitalist systems the essential feature is the extent to which regulation of enterprises is cooperative or confrontational. Policies formulated through a cooperative process that incorporates concerns of the affected actors will engender greater political support and reduce enforcement cost. Moreover, the regulated firms possess much of the technical information needed for effective CAC regulation. Cooperative relations between the firms and the environmental authority facilitate the information processing that is necessary for appropriately articulated CAC policies. The cost saving under market control of pollution in comparison to CAC thus hinges on the nature of negotiations between environmental authorities and the firms they regulate in a CAC framework.
Such inclusion is characteristic of Dutch, French, German, and Japanese environmental policies. In contrast to European traditions of corporatism and the Confucian familial tradition in Japan, regulation in the United States is rather confrontational. Reasons for these differences are rooted in the overall philosophy of how government and private economic activity are related. More cooperative environmental regulatory processes are found in countries where the government generally plays a larger role in economic affairs. The United States is at one extreme, emphasizing lack of governmental involvement. The French and Japanese systems, on the other hand, involve extensive consultation with firms that are regulated, in keeping with the French tradition of indicative economic planning and Japanese industrial development policy. The German and Dutch systems lie in between. This provides an explanation for why market-based regulation has been embraced more fully in the United States in comparison to Europe and Japan. The advantages are greater.
Many applications of market-type instruments for pollution control have been established over the past three decades in the countries considered. However, where market instruments have been introduced, they often are different from the systems conceived in economic theory, being primarily used to supplement direct regulation. Evolution of pollution control policy has been toward hybrid forms, suggesting, as Hahn (1989) has, that the relevant choice is not between the two approaches, but rather one of the correct mix of market and CAC instruments. European countries have established emission charges for a variety of pollutants, and Japan utilized a charge on S[O.sub.2] emissions for many years before phasing it out in 1988. All of these policies were developed to generate revenue for spending on environmental programs rather than for their incentive impacts. Subsidies have not been used as market instruments, but are applied widely to address equity impacts of environmental regulation. Only the United States has established significant tradable permit policies, and these have followed a tentative development, often being hampered by extensive control of the trades. The systems introduced in the United States for S[O.sub.2] and for a variety of air pollutants in the Los Angeles airshed are the most progressive market approaches. In sum, although several market instruments have been applied, true marketization of pollution control policy is quite limited.
Given the time the policy has had to evolve, this observation suggests that the advantages of a market approach over CAC policies are not all that great. On the one hand, market-type instruments lose a degree of cost-effectiveness owing to the ecological complexity of pollution and to political opposition. The full cost-effectiveness of tradable permit and charge systems is attainable only with very complicated structures, and complexity imposes administrative and information costs. Ethics may play a role, but political opposition arises primarily because of the perceived distributional impacts of market instruments. Where market approaches have been used, their structure avoids placing the large burden on firms implicit in an incentive payment high enough to reach environmental quality goals.
On the other hand, CAC policies have the potential to achieve a high degree of cost-effectiveness, but this depends on a highly articulated and flexible use of this approach. The nature of environmental regulation makes the characteristics of a country's administrative practices relevant to achieving this cost-effectiveness. The requirement of a diversified set of directives implies a need to process a great deal of information about the economy. Countries with superior capacity for this can employ CAC approaches more effectively. This capacity, in turn, is enhanced by cooperative interaction between government and regulated firms and the institutions and experience of state intervention in economic affairs. The evolution of market approaches evident in this limited survey is thus quite logical. The United States, having the least cooperative interactions and the least experience with economic planning, has embraced the market approach more fully. The regulatory traditions of the European countries and Japan support them in making CAC cost-effective, which in turn implies that market instruments are more compelling in the United States.
The experience of these five countries with market-type instruments suggests a number of points to be considered by environmental policymakers in developing countries and transition economies. The conceptual discussion points to a handful of national characteristics that are relevant in choosing between market and command strategies. One is the nature of the main sources of pollution, which is tied intimately to information processing: Are there a few large point sources, or are they small and large in number? How diverse are they regarding appropriate control technologies? Fewer and more homogeneous sources favor a command approach. The size of an environmental authority's jurisdiction determines this characteristic to a large extent, which in turn is a function of the size of the country and the extent of centralization of the environmental authority. A second feature is the culture of regulation in the country. The extent of market orientation and the nature of relations between production enterprises and regulators are the key aspects of this dimension. Finally, the administrative capacity of a country for environmental regulation is an important determinant of an appropriate policy strategy. In this regard the level of environmental expertise and extent of corruption(12) within public agencies are key factors, with a wide variance across countries.
In transition countries the relative lack of market culture and deep familiarity with state intervention in the economy favor a command approach. On the other hand the lack of administrative capacity in environmental regulation in some of these countries (Opschoor et al., 1994; Peterson, 1995; Porfiriev, 1997) and their large size and tendency toward centralization argue for market approaches. Although each environmental problem requires individual consideration, a generalization we may draw is that market approaches are more appropriate for China and Russia, while command policies would be cost-effective in countries like Hungary and the Czech Republic. The developing countries are a more diverse group, and generalizations for them are more tenuous. Most of these countries have a market basis for resource allocation and limited experience with state regulation of enterprises. Coupled with their lack of technical expertise in pollution control and possibly more extensive corruption, these features suggest that market approaches may be more cost-effective for them. On the other hand, the relative lack of industrialization and thus fewer sources of pollutants to regulate argue in favor of a command approach.
Although there may be good reasons why regulation of pollution via markets is quite limited, there clearly is an important role for them to play, particularly as complements to policies of direct regulation. The CAC policies to limit lead in gasoline, coupled with tax differentials in Europe and the lead trading program in the United States, are examples of effective hybrid approaches. Surveys (Opschoor et al., 1994; Opschoor & Vos, 1989) reveal that such hybrid policies are very common, and this suggests that the future evolution of environmental regulation well may embrace greater marketization without abandoning a CAC foundation. In the long term, where marketization can offer significant gains in cost-effectiveness, political opposition will be overcome as ways to share the increased efficiency are found.
The author gratefully acknowledges critical input from Paul Burkett, Aimin Chen, Marvin Fischbaum, Donald Richards, and several anonymous referees who read earlier versions of this paper. Thanks are extended to these individuals.
1 See Opschoor & Vos, 1989; and Opschoor et al., 1994, for surveys.
2 Subsidy policies are somewhat different, since acquisition occurs via institution of the policy rather than by a transaction Still, the right is recognized explicitly as acquired legitimately.
3 This was analyzed in a number of articles in the 1970s that were based on the seminal paper by Weitzman (1974). See Watson & Ridker (1984) for the most general treatment. Roberts & Spence (1976) proposed a hybrid permit-charge-subsidy system that theoretically improves efficiency under uncertainty compared to pure systems of either type.
4 The "polluter pays" principle was adopted in 1972 by the OECD as a guiding regulatory philosophy (Organization for Economic Cooperation and Development, 1975). The main motivation was to prevent manipulation of tradable goods prices through environmental policy. More recently the principle has become significant as an ethical premise as well.
5 Figures in current dollars based on purchasing power parity.
6 Feldman (1989) presents a substantially higher share for France, amounting to 3.4% of gross national product (GNP), and Webber (1989) writes that the total cost of regulation in the United States runs to around 2% of GNP The reasons for these discrepancies are not clear.
7 Incomplete data show that this share remained relatively constant through the 1980s. A reviewer has suggested that it has risen to around 2% to 2.5% in recent years.
8 Brown & Johnson (1984) are ambiguous on this point. It is not clear that the cost saving is due to the charge component of the German system, since the diversification of standards across industries may be its source.
9 These are the Minamata (mercury poisoning) and itai itai (cadmium poisoning) diseases.
10 Another part of the air pollution compensation fund (around 20%) was provided by the automobile weight tax (Marcus, 1989).
11 Sample size was 15. When regions with the most and least abatement were dropped (n = 13), the correlation coefficient rose to .92 (Bressers, 1988, p. 509).
12 The implications of corruption for instrument choice are not clear. Oosterhuis and de Savomin Lohman (1994) argue that direct regulation offers greater scope for bureaucratic discretion than does environmental taxation. Prima facie, this suggests a reduction in the effectiveness of CAC policies vis-a-vis emission charges where corruption is a feature.
Ayres, R. U., & Kneese, A. V. (1969). Production, consumption and externalities. American Economic Review, 59(3), 282-297.
Baumol, W., & Oates, W. (1979). Economics, environmental policy, and the quality of life. Englewood Cliffs, NJ: Prentice-Hall.
Baumol, W., & Oates, W. (1988). The theory of environmental policy. Cambridge: Cambridge University Press.
Bohm, P., & Russell, C. (1985). Comparative analysis of alternative policy instruments. In A. Kneese & J. Sweeney (Eds.), Handbook of natural resources and energy economics, vol. 1 (pp. 397-455). Amsterdam: North-Holland.
Braden, J. B., & Kolstad, C. D. (Eds.) (1991). Measuring the demand for environmental quality. Amsterdam: North-Holland.
Bressers, J. T. A. (1988). A comparison of the effectiveness of incentives and directives: The case of the Dutch water quality policy. Policy Studies Review, 7(3), 500-518.
Brown, G. M., & Johnson, R. W. (1984). Pollution control by effluent charges: It works in the Federal Republic of Germany, why not in the U.S.? Natural Resources Journal, 24 (4), 929-966.
Buchanan, J. M., & Tullock, G. (1975). Polluters' profits and political response: Direct controls versus taxes. American Economic Review, 65(1), 139-147.
Callan, S. J., & Thomas, J. M. (1996). Environmental economics and management: Theory policy and applications. Chicago, IL: Irwin.
Coase, R. (1960). The problem of social cost. Journal of Law and Economics, 3, 1-44.
Cropper, M., & Oates, W. (1992). Environmental economics: A survey. Journal of Economic Literature, 30(2), 675-740.
Cushman, J. H. (1991, October 30). U.S. proposes regulations to decrease acid rain. New York Times, pp. A1, A10.
Cushman, J. H. (1995, March 17). Proposed changes simplify rules on pollution control. New York Times, p. C1.
Dales, J. H. (1968). Pollution, property and prices: An essay in policy-making and economics. Toronto: University of Toronto Press.
de Savomin Lohman, L. (1994). Economic incentives in environmental policy: Why are they white ravens? In J. B. Opschoor & R. K. Turner (Eds.), Economic incentives and environmental policies: Principles and practice (pp. 120-138). Dordrecht: Kluwer Academic.
Dudek, D. J., & Palmisano, J. (1988). Emissions trading: Why is this thoroughbred hobbled? Columbia Journal of Environmental Law, 13 (2), 217-256.
Feldman, D. L. (1989). France. In F. N. Bolotin (Ed.), International public policy sourcebook: volume 2, Education and environment (pp. 235-255). New York, NY: Greenwood Press.
Freeman, A. M. (1993). The measurement of environmental and resource values: Theory and methods. Washington, DC: Resources for the Future.
Hahn, R. W. (1989). Economic prescriptions for environmental problems: How the patient followed the doctor's orders. Journal of Economic Perspectives, 3(2), 95-114.
Hahn, R. W., & Hester, G. L. (1989). Marketable permits: Lessons for theory and practice. Ecology Law Quarterly, 16(2), 361-406.
Hayek, F. A. (1945). The use of knowledge in society. American Economic Review, 35 (4), 519-530.
Hockenstein, J. B., Stavins, R. N., & Whitehead, B. W. (1997). Crafting the next generation of market-based environmental tools. Environment, 39(4), 13-20, 30-33.
Huppes, G., & Kagan, R. A. (1989). Market-oriented regulation of environmental problems in the Netherlands. Law and Policy, 11 (2), 215-239.
International Monetary Fund. (1995). Government finance statistics yearbook. Washington, DC: International Monetary Fund.
Johansson, P. (1987). The economic theory and measurement of environmental benefits. Cambridge: Cambridge University Press.
Kelman, S. (1981). What price incentives? Economists and the environment. Boston, MA: Auburn House.
Keohane, N. O., Revesz, R. L., & Stavins, R. N. (1997, January). The positive political economy of instrument choice in environmental policy. Conference paper for Allied Social Science Associations, New Orleans.
Kolstad, C. (1987). Uniformity vs. differentiation in regulating externalities. Journal of Environmental Economics and Management, 14 (4), 386-399.
Kopp, R. J., Portney, P., & DeWitt, D. E. (1990). International comparisons of environmental regulation. In Environmental policy and the cost of capital (pp. 75-91). Washington, DC: American Council for Capital Formation.
Lesser, J. A, Dodds, D. E., & Zerbe, R. O. (1997). Environmental economics and policy. Reading, MA: Addison-Wesley.
Liroff, R. A. (1986). Reforming air pollution regulation: The toil and trouble of EPA's bubble. Washington, DC: The Conservation Foundation.
Marcus, A. A. (1989). Japan. In F. N. Bolotin (Ed.), International public policy sourcebook: volume 2, Education and environment (pp. 275-291). New York, NY: Greenwood Press.
Miller, L. (1989). Federal Republic of Germany. In F. N. Bolotin (Ed.), International public policy sourcebook: volume 2 Education and environment (pp. 207-234). New York, NY: Greenwood Press.
Miyamoto, K. (1991). Japan in European environmental yearbook (pp. 665-676). London: International Institute for Environmental Studies.
Montgomery, W. D. (1972). Markets in licenses and efficient pollution control programs. Journal of Economic Theory, 5 (3), 395-418.
Oates, W. E. (1994). Environment and taxation: The case of the United States. In Organization for Economic Cooperation and Development (Ed.), Environment and taxation: The cases of the Netherlands, Sweden and the United States (pp. 112-160). Paris: Organization for Economic Cooperation and Development.
Oosterhuis, F. H., & de Savomin Lohman, A. F. (1994). Environment and taxation: The case of the Netherlands. In Organization for Economic Cooperation and Development (Ed.), Environment and taxation The cases of the Netherlands, Sweden and the United States (pp. 3-40). Paris: Organization for Economic Cooperation and Development.
Opschoor, J. B., de Savomin Lohman, A. F., & Vos, H. B. (1994). Managing the environment: The role of economic instruments. Paris: Organization for Economic Cooperation and Development.
Opschoor, J. B., & Vos, H. B. (1989). Economic instruments for environmental protection. Paris: Organization for Economic Cooperation and Development.
Organization for Economic Cooperation and Development. (1975). The polluter pays principle: Definition, analysis, implementation. Paris: Organization for Economic Cooperation and Development.
Organization for Economic Cooperation and Development. (1993). OECD environmental data compendium 1993. Paris Organization for Economic Cooperation and Development.
Organization for Economic Cooperation and Development. (1994a). The distributive effects of economic instruments for environmental policy. Paris: Organization for Economic Cooperation and Development.
Organization for Economic Cooperation and Development. (1994b). Environment and taxation: The cases of the Netherlands Sweden and the United States. Paris: Organization for Economic Cooperation and Development.
Organization for Economic Cooperation and Development. (1997). OECD national accounts, main aggregates, vol. 1. Paris: Organization for Economic Cooperation and Development.
Peterson, D. J. (1995). Building bureaucratic capacity in Russia: Federal and regional responses to the post-Soviet environmental challenge. In J. DeBardeleben & J. Hannigan (Eds.), Environmental security and quality after communism: Eastern Europe and the Soviet successor states (pp. 105-125). Boulder, CO: Westview.
Pigou, A. C. (1938). The economics of welfare (4th ed.). London: Macmillan.
Porfiriev, B. (1997). Environmental policy in Russia: Economic, legal and organizational issues. Environmental Management, 21 (2), 147-157.
Roberts, M. J., & Spence, M. (1976). Effluent charges and licenses under uncertainty. Journal of Public Economics, 5, 193-208.
Ruff, L. (1970). The economic common sense of pollution. The Public Interest, 19, 69-85.
Stiglitz, J. (1994). Whither socialism? Cambridge, MA: MIT Press.
Tietenberg, T. (1978). Spatially differentiated air pollutant emission charges: An economic and legal analysis. Land Economics, 54(3), 265-277.
Tietenberg, T. (1996). Environmental and natural resource economics (4th ed.). New York, NY: HarperCollins.
Totsuka, T. (1989). Japan. In E. J. Kormandy (Ed.), International handbook of pollution control (pp. 323-336). New York, NY: Greenwood Press.
Turner, K., & Opschoor, H. (1994). Environmental economics and environmental policy instruments: Introduction and overview. In J. B. Opschoor & R. K. Turner (Eds.), Economic incentives and environmental policies: Principles and practice (pp. 3-42). Dordrecht: Kluwer Academic.
Vogel, D. (1986). National styles of regulation: Environmental policy in Great Britain and the United States. Ithaca, NY: Cornell University Press.
Watson, W. D., & Ridker, R. G. (1984). Losses from effluent taxes and quotas under uncertainty. Journal of Environmental Economics and Management, 11(4), 310-326.
Webber, D. J. (1989). United States. In F. N. Bolotin (Ed.), International public policy sourcebook: volume 2, Education and environment (pp. 335-352). New York, NY: Greenwood Press.
Weitzman, M. (1974). Prices vs. quantifies. Review of Economic Studies, 41, 447-491.
World Resources Institute. (1994). World resources 1994-95. Oxford: Oxford University Press.
Richard Lotspeich received degrees in economics from Georgetown University and the University of New Mexico. After a fellowship at the Los Alamos National Laboratory, he undertook postdoctoral studies on the Soviet Union at Indiana University. Currently he is an associate professor at Indiana State University, where his work focuses on environmental economics and transition economies.…