Environmental Policy Instruments for Dematerialisation of the ...

Environmental Policy Instruments for Dematerialisation of the European Union

Arno Behrens

No. 7, May 2004 ISSN 1729-3545

Sustainable Europe Research Institute (SERI) Garnisongasse 7/27, 1090 Vienna, Austria Tel.: +43-1-9690728-0, Fax: +43-1-9690728-17 www.seri.at, [email protected]

The author Arno Behrens holds an M.A. in economics and is currently researcher at the Sustainable Europe Research Institute (SERI) in Vienna, Austria.

Contact: Arno Behrens Sustainable Europe Research Institute (SERI) Garnisongasse 7/27 A – 1090 Vienna, Austria E-mail: [email protected]

An extended version of this paper is available at www.seri.at/Data/projects/students/downloads/MasterThesis_ArnoBehrens_2004.pdf

SERI Background Papers present a comprehensive overview on the state of the art in one of SERI's research areas, addressing researchers from related research fields.

2

Abstract This paper deals with environmental policy instruments designed to reduce the material input of European economies. In line with sustainable development, these instruments need to allow for economic growth, while taking into account environmental and social considerations. With regard to the environmental dimension, dematerialisation has been pointed out as one of the main goals, recognising that industrialised economies are capturing an unsustainably large share of world resource use, the consequences of which are faced with considerable degrees of scientific uncertainty. Focussing on the inputs of the economic metabolism will also lead to improvements regarding its outputs, due to the fact that all inputs are ultimately turned into waste, emissions and other outputs. Following some introductory paragraphs on dematerialisation and policy instruments in general, the main part of this paper deals with specific instruments for policy-makers to pursue a dematerialisation strategy. As a start, instruments that support voluntary behavioural changes are presented. They include the design of a clear concept of dematerialisation that needs to be publicly communicated; the provision of information to all involved actors on how to comply with that concept; the development of adequate resource-management systems in order to reduce material requirements of products over their whole life-cycle; as well as the provision of environmentally relevant education on all levels. This section also features ownership rights and cooperation between actors as well as an extensive part on voluntary environmental agreements. The second section deals with economic instruments and the incentives for economic agents to reduce material input associated with them. Subsidies, taxes, certificates, and public procurement are discussed in this section. The third section covers some aspects of regulatory policies. It is followed by some concluding remarks, stressing the need for an appropriate mix of instruments in order to put the concept of dematerialisation into practice.

3

1.

Introduction

The past decades have been characterised by an increase in European societies’ environmental consciousness. Reasons for this include increased resource use and rising pollution levels caused by the economic system on the one hand, but also better knowledge about the dangers of environmental damage to human kind on the other. This development has affected policy-makers, who have been facing an ever higher pressure to protect natural resources and who are now recognising the importance of environmental policy as an important point on their political agenda. However, the enforceability of environmental policy measures still depends on the situative conditions for action. The visibility and topicality of certain environmental problems simplifies the implementation of corrective measures. However, these short-run oriented measures often dwarf long-term environmental problems, whose threats are either underestimated or simply not realised. The use of natural resources has been described in this context, with the consequences of their excessive use lying far beyond the scope of policy-makers. In addition, social preferences shift in times of economic crises at the expense of an ambitious environmental policy agenda. Such trends need to be counteracted by raising environmental questions as well as by the use of innovative policy instruments. It shows that environmental policy is a tightrope walk between ecological necessity and political enforceability, the latter of which depends largely on the choice and application of the right policy instruments. 1.1

Dematerialisation and Sustainable Development

In an effort to encounter the environmental challenges posed by an ever increasing scale of the economic subsystem, a preventive strategy has been formulated by the proponents of dematerialisation. This strategy recognises that it is not only the environmental pressure of specific pollutants, but also the enormous amounts of energy and material flow inputs to the economic subsystem that pose the central ecological problem (Nachhaltigkeit, 2001). It thus calls for reducing all material flows in order to avoid potential environmental risks – especially in view of the environmental uncertainties associated with most economic activities, the dangers associated with irreversible processes, and the assumption that any movement of materials can lead to changes in the conditions of the ecosystem and thus to an increase of the risk of ecologically undesirable developments (Hans Böckler Foundation, 2000). Recognising that “any flows of material into the economy will lead to output flows sooner or later, many of them at other locations and with changed composition” (EEA, 2002: 101), dematerialisation requires a considerable reduction of land, energy and materials consumption in order to be sufficient. The advantages of focussing on restricting input into the economy in the long run over limiting output for reducing generic pressure on the environment are twofold. First, input oriented environmental policies act on the cause of ecological problems 4

rather than on the symptoms. The reduction of inputs reduces potential consequences of economic activity on the environment and thus potential external effects (Stewen, 2002). Second, input oriented environmental policies are regarded to be more efficient than conventional environmental policy measures, due to the fact that the quantity of materials to be used as inputs is relatively well known and due to new incentives for the development of more resource efficient technologies (Ibid.). In general, input orientation is believed to be better suited for dealing with the complex links among population, poverty, growth, resources and the environment. Dematerialisation can either be absolute or relative. Absolute dematerialisation, also referred to as strong dematerialisation, occurs where total material input of an economy decreases in absolute terms. Relative dematerialisation, or weak dematerialisation, refers to an increase in the intensity of use, requiring the ratio between material input and GDP to fall over time (Moll et al., 2003). Unlike absolute dematerialisation, relative dematerialisation is associated with an increase in resource use, however, at a slower pace than economic growth. Both of these concepts necessitate a “break in the link between ‘environmental bads’ and ‘economic goods’” (OECD, 2002: 4). This is referred to as de-linking or decoupling economic from the use of environmental resources. As with dematerialisation, decoupling can also be either absolute or relative. Absolute decoupling occurs where “the environmentally relevant variable is stable or decreasing while the economic driving force is growing” (Ibid.). Relative decoupling, on the other hand, refers to the growth rates of the two, taking place where “the growth rate of the environmentally relevant variable is positive, but less than the growth rate of the economic variable” (Ibid.). Most dematerialisation oriented concepts, such as the ones brought forward by the factor X debate, call for absolute de-linking, with the objective to increase welfare without increasing the use of nature in general and the use of material inputs in particular. The necessity of a reduction of material flows not only for environmental reasons but also for social ones regarding intragenerational and intergenerational fairness has been recognised by the World Commission on Environment and Development in 1987. Its definition of sustainability is based on two key concepts: needs and limits. The concept of needs refers to the social dimension of sustainable development and the commitment to eradicate poverty by “meeting the basic needs of all and extending to all the opportunity to satisfy their aspirations for a better life” (WCED, 1987: 44). Regarding dematerialisation, this concept bears recognition of the fact that developing countries should have the opportunity and the ability to achieve economic growth. Taking into account that economic growth will require an increase in the use of energy and environmental resources (at least up to a certain level of development) in those nations, industrialised countries will be required to considerably reduce energy and material flows in order for the planet’s resource and sink functions to remain intact. Evidence of the reduction requirements in industrialised countries is given by the fact that per capita energy use in industrialised countries is ten times higher than the one in developing countries (Hinterberger et al., 5

1996), and the fact that 80 percent of worldwide resource flows accrue to 20 percent of the world’s population in industrialised countries (Ibid.). The ‘factor X’ debate gives suggestions for reduction targets, reaching from at least four (i.e. a fourfold increase in resource productivity) to 10 and more within the next 30 to 50 years (cp. Nachhaltigkeit, 2001). The need to meet ambitious dematerialisation targets is emphasised by the second concept. The concept of limits “recognises that the current state of technology and social organisation imposes limits on the ability of the environment to meet present and future needs, so we must moderate our demands on the natural environment” (Carter, 2001). A reduction of inputs in industrialised nations and a focus on less material- and energy-intensive growth in all countries will contribute to remaining within the limits posed by the environment. In view of the ever increasing scale of the economic subsystem, governments need to commit themselves to sustainable development as the only alternative for the future in order for dematerialisation to become reality. The recognition of economic, ecological and social goals as equally important policy objectives needs to be the basis for achieving economic growth while minimising environmental impacts and social inequalities. 1.2

Dematerialisation and Policy

In the context of this paper, the definition used for environmental policy instruments calls for “structured activities aimed at changing other activities in society towards environmental goals” (Huppes, 2000: 5). To be more precise, the economic system needs to be influenced in such a way “that the environmental impact of the material flow caused by those systems is significantly […] reduced” (te Riele, 2001: 13). Hence, with dematerialisation conceptualised as the primary environmental policy goal in this paper, instruments need to be assessed for their ability to reduce resource flows related to the economic system. These include all movements of materials actively caused by human beings with the distinction between input, throughput and output (Stewen, 2000). The need to comply with the conceptual design of the general regulatory framework applies to environmental policy as much as to any other government policy area. The selection and design of the instruments is thus of substantial significance. Criteria for such processes have been laid out by the OECD and the Rio Declaration (1992), stressing the superiority of economic (market oriented) measures over direct regulations, commonly referred to as command-and-control mechanisms (Altmann, 1997). The latter consist of strategies aiming at regulating activities that harm the environment based mainly on legislative and regulatory provisions (IISD, 2003; James, 1997). Examples include bans for using certain resources, input quotas, restrictions on emissions, ambient water quality standards etc. They are enforced by means of litigation, licensing and fines or other penalties in cases of noncompliance (IISD, 2003). 6

Economic instruments, on the other hand, work by using market forces and financial incentives. They are intended to give incentives for environmentally friendly actions by means of rewarding them while penalising activities that are harmful to the environment, creating a direct impact on cost/benefit decisions in favour of environmentally friendly behaviour. As such, they “usually allow for adaptive choice and decentralised decision-making by those whose behaviour is to be modified” (James, 1997: 12). It will be shown later that they thus increase economic efficiency and that they allow for attaining standards higher than those imposed by regulations. Additionally, they reduce costs for enforcement and administration. It needs to be stressed, however, that the border between the two types of policy instruments is often not very clear, since economic instruments require an appropriate regulatory basis to be effective. “Wherever economic instruments have been used […] supporting regulations have been applied” (James, 1997: 12). Despite the increasing importance of economic instruments, regulative measures are still dominating in practice. One possible explanation for that is the fact that markets alone will not contribute significantly to more sustainable resource use. On the contrary, voluntary internalisation of external costs seems rather unlikely (Altmann, 1997). This justifies government regulation for sustainable development by means laid out below. The approach chosen for the presentation of the instruments in this paper follows the ones used by Jänicke et al. (1999) and Hinterberger et al. (1996) regarding the intensity of government intervention and therefore the intensity by which they affect individual decision-making. As a start, instruments that support voluntary behavioural changes will be presented. They include the design of a clear concept of dematerialisation that needs to be publicly communicated; the provision of information on how to comply with that concept to all actors involved; the development of adequate resource-management systems in order to reduce material requirements of products over their whole lifecycle; as well as the provision of environmentally relevant education on all levels. This part also features ownership rights and cooperation between actors as well as an extensive part on voluntary environmental agreements. Chapter 3 deals with economic instruments and the incentives associated with them. Subsidies, taxes, certificates, and public procurement will be discussed in that section. The fourth chapter will cover some aspects of regulatory policies. It is followed by some concluding remarks, stressing the need for an appropriate mix of instruments in order to put the concept of dematerialisation into practice.

2.

Voluntary Instruments

Of all policy instruments, those supporting voluntary changes towards more sustainable resource use cause the least interference with individual freedom. Four different areas of policy intervention have been pointed out by Hinterberger et al. (1996). First, an institutionalised concept needs to be formulated and appropriately 7

communicated to the public in order to create consensus on the goals of environmental policy. Second, all relevant actors need sufficient information on how to behave in accordance with the above concept, i.e. on how to implement it in practice. The third area deals with another aspect of information: environmental information and management systems in general and eco-audits specifically, as promoted by the European Union, need to be developed into integrated resource management systems with the aim of reducing the material intensity of all goods and services. The fourth main area of intervention mentioned is environmental education. Following the presentation of those four areas of intervention, there will be a discussion on voluntary approaches as alternatives to legislation for businesses and industries. 2.1

Establishing the Concept of Dematerialisation

A pre-requisite for policy measures focussing on voluntary actions is a clear concept of dematerialisation (e.g. Factor 10), providing “ecological guard-rails for long term economic development” (Schütz/Welfens, 2000). Such a concept lays out “common rules for decision making understandable and relevant for all” (Schütz/Welfens, 2000), creating and facilitating certain behavioural options when well communicated to the public. It should be allowed to evolve in societal processes resulting in institutional change. Policy measures can therefore only try to influence such a process. This can be done through the media by supporting the publication of new insights or views, by initiating study commissions where scientists and politicians come together for public discussions, by initiating research projects with the results published and made freely available etc. (Hinterberger et al., 1996), to name but a few examples. The result of such ‘thought-provoking impulses’ is a publicly held debate on dematerialisation leading to slow institutional changes in society associated with a change of preferences, thus leading the economic units to cut environmental use voluntarily by means of sustainable production and consumption patterns. 2.2

Raising Information and Awareness

Another imperative for sustainable development is sufficient environment related information on the macro as well as on the micro level. More and better information is needed about the threats posed by an ever increasing scale of the economic metabolism. In fact, “unless firms and consumers are well informed, they may take actions that are not in their own interests. Unless decisions are based on good information, markets will not work well. But in a modern, complex, free market economy, firms and consumers are unlikely to be well informed about the consequences of all their decisions” (Begg et al., 1994: 52). This also applies to policy-makers, who need environmental data for successful planning and decisionmaking. Without such information, their decisions will be “little better than best guesses and are likely to be wrong” (UNEP, 2002a: 404). However, “high quality, 8

comprehensive and timely information on the environment remains a scarce resource” (Ibid.). Reasons for this include the difficulty to obtain the right data, and the problem of finding indicators that “capture and reflect the complexity of the environment and human vulnerability to environmental change” (Ibid.). Even though it needs to be accepted that the consequences of certain environmental damages cannot be known, pertinent, sound, accurate and reliable information can help to place decisions on a solid basis with more calculable outcomes. However, they do not solve the problem for consumers and producers to make the economic and environmentally right decision within a certain policy framework. For consumers to behave in an environmentally friendly way requires them to know about the environmental impact of the production, use and disposal of a good or service. Such information can serve as an environmental criterion for buying or not buying a product. In cases of high environmental awareness, this increases pressure on producers of products with high environmental impacts to rethink their strategy. Unfortunately private markets tend not to produce such information because producers have little incentive to do so. Even though many products are branded as ‘ecological’ or ‘biological’ and marked with labels that ‘prove’ their environmental friendliness, there is often no indication of the total material or energy use over their whole life cycle. Instead of focussing on details, an integrated, comparable and easily understandable labelling system should be introduced. As a start, this could focus on the material intensity of a product ‘from cradle to grave’ using the material intensity analysis (MAIA) methodology, for example. The problem of imported goods and services that have not been evaluated for their material use can be solved temporarily by using average values or approximations until the relevant environmental management and accounting systems have been installed abroad (Hinterberger et al., 1996). 2.3

Environmental Information and Management Systems

Another possibility for publishing environmentally relevant information is the eco-audit as regulated by the Council Regulation (EEC) No. 1836/93 of 29 June 1993, also known as EMAS (Environmental Management and Auditing Scheme) Regulation, “allowing voluntary participation by companies in the industrial sector in a Community eco-management and audit scheme” (EEC, 1993: 1). It builds on the possibility for companies in the industrial sector to voluntarily exceed the legal provisions in force and hence to make further improvements in their internal environmental protection arrangements. Companies signing up for the scheme will become objectively judged by independent internal or external experts by means of an environmental audit. This refers to a “management tool comprising a systematic, documented, periodic and objective evaluation of the performance of the organisation, management system and processes designed to protect the environment with the aim of: (i) facilitating

9

management control of practices which may have impact on the environment; (ii) assessing compliance with company environmental policies” (EEC, 1993: 2 et seqq.). Following its validation and certification the company is allowed to use the EMAS symbol in its external communications, showing its dedication to environmental protection with a clear sign to its costumers, regulators and investors. One of the drawbacks of the eco audit is the fact that it aims at production sites and leaves out the products themselves (Hinterberger et al., 1996). This implies that companies whose production sites fulfil all relevant EU regulations but whose products have a high impact on the environment or are very material intensive can still be awarded the certificates associated with the EMAS Regulation. On the other hand, it is also possible for companies to adopt dematerialisation as their corporate environmental goal for specific production sites. Such a goal cannot be reached solely by changing production processes. The corporate environmental policy should also focus on the selection of preliminary products according to their material and energy efficiency and on decreasing the material intensity of the products, e.g. by increasing their usability. In all, the environmental goal should be the reduction of the material intensity per service unit (MIPS) (Hinterberger et al., 1996). The quantity of such a voluntary achievement will depend on the possibilities of different regions, industries, production sites, products etc., but could be positively influenced by a well communicated guiding concept of dematerialisation. 2.4

Environmental Education

Another priority on the policy agenda should be the stimulation of environmental education and knowledge with the aim of increasing public awareness and knowledge about environmental problems. The focus should lie in providing the public “with the necessary skills to make informed decisions and take responsible action” by teaching individuals “how to weigh various sides of an issue through critical thinking” and by enhancing “their own problem-solving and decision-making skills” (EPA, 2003). A holistic, interdisciplinary approach should be stressed resulting in “an overall perspective which acknowledges the fact that natural environment and man-made environment are profoundly interdependent” and that “the acts of today [are linked] to the consequences of tomorrow” (UNESCO-UNEP, 1978: 2). Five objectives of environmental education have been highlighted by the world’s first Intergovernmental Conference on Environmental Education held in Tbilisi, Georgia in 1977, including awareness concerning the environment and its problems; knowledge concerning experience in, and basic understanding of the environment; attitudes regarding a set of values and feelings for concern for the environment; skills for identifying and solving environmental problems; and participation regarding active involvement in working towards resolving environmental problems (UNESCO-UNEP, 1978: 3).

10

2.5

Material Intensity and Property Rights

The allocation of property rights can be an appropriate tool to internalise external costs. One rather new aspect of it is referred to as extended producer responsibility (EPR). It is defined as “an environmental policy approach in which a producer’s responsibility, physical and/or financial, for a product is extended to the postconsumer stage of a product’s life cycle. There are two related features of EPR policy: (1) the shifting of responsibility (physically and/or economically; fully or partially) upstream to the producer and away from municipalities, and (2) to provide incentives to producers to incorporate environmental considerations in the design of their products” (OECD, 2001a: 18). This approach bears recognition to the fact that producers have a “considerable ability to reduce the life-cycle impacts of their products” (EPA, 1998a: 1). However, numerous other actors along the product chain need also to be involved on order to achieve better resource use and pollution prevention. This view is represented by the notion of extended product responsibility. It calls for “lasting and substantial environmental improvements in product systems” (Ibid.) with the “combined expertise, ingenuity, cooperation and commitment” (Ibid.) of suppliers, designers, manufacturers, distributors, retailers, customers, recyclers, remanufacturers, and disposers. For example, consumers play a critical role in selecting products that are characterised by low material intensity. However, this approach also highlights the producers’ “unique position - through their capability to affect product design, material choices, manufacturing processes, and product delivery - to reduce the lifecycle environmental impacts of their products” (EPA, 1998b: 2). In other words, both approaches described above are bearing recognition to the polluter-pays principle, i.e. by internalising social costs “within the product chain responsible for generating the externality” (OECD, 2001a: 21). Industry-based initiatives are increasingly being adopted especially in the fields of product stewardship and take-back. Other widely used voluntary strategies are leasing and ‘servicising’. The latter refers to firms evolving into service providers, selling “their costumers function rather than physical input” (OECD, 2001a: 45). This creates incentives for producers to increase the economic life-time and to simplify maintenance of their products, thus allowing for new strategies to minimise resource consumption while maximising resource productivity. Another strategy consistent with EPR is for firms to voluntarily increase the warranty period for their products. With increased acceptance of the concept of dematerialisation, a competition about warranty periods could emerge, providing an effective indication for product’s expected lifetime (Hinterberger et al., 1996). 2.6 Voluntary Environmental Agreements of Businesses or Industries Various papers have been published on the effectiveness and efficiency of voluntary agreements, and there seems to be little dispute over the fact that firms can profit 11

from taking voluntary action. Their usefulness for policy-makers to achieve environmental targets, however, is less clear. On one hand, it is argued that voluntary approaches offer “a chance to address environmental problems in a flexible manner at a low cost, based on consensus-building between the different stakeholders” (OECD, 2003: 10). Critics on the other hand, believe that “such approaches provide few environmental improvements beyond what would have occurred anyway, while both administrative and abatement costs could be greater than using other instruments” (Ibid.). In its 1999 assessment of voluntary approaches, the OECD distinguishes between four different ways to implement environmental policy in general and EPR specifically. They are presented in increasing order of importance that public authorities play in their application: industry-based unilateral commitments; private agreements between polluters and pollutees; agreements negotiated between industry and public authorities; and public voluntary programmes (OECD, 1999a). Industry-based unilateral commitments “consist of environmental improvement programmes set up by firms and communicated to their stakeholders” (OECD, 1999a: 16). The compliance with pre-defined environmental targets may be left to be controlled by external parties to increase credibility. Private agreements between polluters and pollutees are defined as “contracts between a firm (or sometimes a group of firms) and those who are harmed by its emissions […] or their representatives” (Ibid.). Such contracts are usually followed by the introduction of environmental management programmes and/or pollution abatement devices. Agreements between industry and public authorities, also in the form of contracts, usually include certain targets to be reached within a defined timeframe. Finally, public voluntary programmes constitute frameworks created by public authorities, leaving the choice to participate to the individual firm. They require “participating firms [to] agree to standards […] which have been developed by public bodies such as environmental agencies” (OECD, 1999a: 18). Terms of membership, rules to be complied with, as well as monitoring and evaluation criteria are set by the programme. Participation can include economic benefits such as research and development subsidies, technical assistance etc. (Ibid.). Even though voluntary approaches may contain ‘trigger clauses’ that cause environmental regulations to be ‘triggered’ in case of non-compliance with the contract (OECD, 2001a), they are widely regarded to constitute effective means for industries to avoid costs associated with government intervention in the form of regulations. This being the ultimate goal of businesses justifies another categorisation based on the purpose of voluntary approaches relative to environmental regulations. They can thus be divided into substitutive, integrative, anticipatory, and applicative agreements. By substituting regulations, voluntary agreements “are able to reach the same objectives with a lower collective cost and/or with a better allocation of costs” (Croci/Pesaro, 1999: 9). With their integration into regulatory measures, they can lead to “better environmental results with a group of 12

pioneer firms due to lower marginal costs for improvements or due to their greater competitive advantages” (Ibid.). Anticipating future regulation, voluntary agreements can help public administration to obtain information on the costs of regulation. Furthermore, they can serve as a means to “gradually predispose the industrial system to the most stringent limits” (Ibid.). At last, voluntary agreements can serve to implement existing law at the European, national or local level. On the European level, the communication adopted by the commission entitled “Environmental Agreements at Community Level Within the Framework of the Action Plan on the Simplification and Improvement of the Regulatory Environment” [COM(2002) 421 final] defines environmental agreements as typically unilateral commitments from businesses or industry designed to achieve environmental objectives such as pollution abatement or material input reductions and distinguishes between three different sources they can come from: purely spontaneous decisions by stakeholders, responses by stakeholders to an expressed intention of the Commission to legislate; or at the initiative of the Commission. However, voluntary agreements are not legally binding on the Community level because the Commission is not authorised to conclude legally binding voluntary agreements (Ökobüro/ÖGUT, 2003). It has therefore developed a framework of two alternatives to acknowledge environmental agreements: self-regulation and coregulation. The former refers to a procedure (usually initiated by the stakeholders) where the Commission acknowledges an agreement either by a simple exchange of letters or by a Commission Recommendation. As such, self-regulation does not require a legislative act. It may be preferable in situations where existing agreements are sufficient to regulate certain issues. For the effective implementation of such agreements, proper monitoring and reporting are regarded as crucial. It may therefore be necessary to complement a Commission Recommendation by a Decision of the European Parliament and the Council to establish an appropriate monitoring system. This allows for the results of an environmental agreement to be monitored more closely. Coregulation, on the other hand, represents a more binding and formal manner of reaching environmental agreements within the framework of a legislative act. The essential aspects of such a framework are established by the legislators, including “the objectives to achieve, the deadlines and mechanisms relating to its implementation, methods of monitoring the application of the legislation and any sanctions which are necessary to guarantee the legal certainty of the legislation” (European Commission, 2002: 8). Environmental agreements will be required to ensure the commitment of economic operators to the implementation of these modalities. Coregulation may thus be regarded as a hybrid between selfregulation and command-and-control measures, intended to create administrational benefits as well as advantages for industries and businesses. It is “usually initiated by the Commission, either on its own initiative or in response to voluntary action on part of industry” (Ibid.). The successful implementation of voluntary approaches requires a set of prerequisites, which have been outlined by Croci and Pesaro (1999): the definition of 13

clear and attainable objectives; the recognition of reciprocal advantages for the parties; the threat of retaliation by public authorities in case of non-compliance (which also requires effective monitoring systems); and trust between the parties. This list can be extended by adding the need for participation of all relevant actors, and transparency of all processes and agreements, as proposed in the communication of the European Commission mentioned above. Even when these conditions are met, a number of disadvantages or problems can still arise. Objections against voluntary approaches other than those mentioned above include the possibility of competition restricting agreements between firms of an industry due to internal agreements on which individual firms can achieve the commitments made in the most cost-effective way (Wicke, 1993). Another one is concerned with the consequences in cases where the objectives of the agreements have not been met. The danger of environmental problems to become ‘eternalised’ in such cases seems especially relevant due to the time-horizon of voluntary agreements being very long in some cases, and retaliatory environmental measures by public authorities taking time to be implemented (Ibid.). Advantages, on the other hand, include the fact that industries or individual companies can choose their own strategies to comply with the contracts. Thus, the most cost-effective way can be selected. Also, voluntary approaches reduce the need for costly and tedious law-making procedures and the disputes and resistance associated with them. One of the most striking advantages of voluntary approaches, however, is their flexibility. They can be implemented on products as well as on production processes and product chains; on the local, regional, national or EU-level; differentiated according to the size of a company; as integrated frameworks or as plans by stages subject to regular control; but also in combination with subsidies, incentives for innovation or other additional provisions (Ibid.). In conclusion, voluntary agreements are recommended to complement legislation rather than to substitute it. By using them in conjunction with commandand-control and/or market-based instruments the risk of insufficient abatement can be reduced. Unfortunately, there are still only few voluntary initiatives that “are directly linked with government policy and regulatory framework in a way that would complement the strengths and weaknesses of both” (UNEP, 2002b: 33)

3.

Economic Instruments

With the development of environmental policy over the past decades, an increasing emphasis has been placed on the design of policy instruments that allow for the internalisation of external environmental costs. Policy-makers “face the challenge of identifying policies and strategies that make it in everyone’s economic interest to utilise environmentally sound products and services” (Huber et al., 1998: 7). Unlike the traditional instruments of environmental policy such as regulations, market-based systems of incentives and disincentives allow policy-makers to achieve environmental objectives in the most cost-effective way. A definition of economic 14

instruments is given by the OECD, describing them as “instruments that affect costs and benefits of alternative actions open to economic agents, with the effect of influencing behaviour in a way which is intended to be favourable to the environment” (OECD, 1991 in AIM, 1997: 7). Generally identified with taxes, subsidies, tradeable permits, user charges etc., they deal with “incentives, flexibility, and understanding people’s behaviour and self-interest” (UNEP, 1994). The principal objective of economic instruments is the reduction of externalities (i.e. environmental degradation costs and resource depletion costs arising due to market, institutional, and policy distortions). The resulting incorporation of social costs into individual cost-benefit calculations is intended to lead to increased cost consciousness and a higher rate of innovation. However, as opposed to regulatory measures, market-oriented instruments, by making use of the price mechanism, “allow polluters and resource users to find their own best mix of controls or responses and therefore result in lower private costs than other approaches” (Huber et al., 1998: 17). As such, economic instruments are very important in bringing about an ecologically oriented structural change. This is also reflected in their increased role in policy design resulting not only in a broader application amongst OECD countries, but also in a greater variety of instruments applied within the context of economic instruments (OECD, 1999b). The focus for the use of this paper is placed on the possible influence of market based instruments on material intensity. Those with special relevance include subsidies, taxes and certificates. They will further be investigated as follows. 3.1

Reform of the Subsidy System

Reforming or abolishing subsides has been ranking very high on the European political agenda for over a decade. The need for such measures is undisputed, however, the discussion focuses on where and how to restructure the system. With regards to improving the environment in general and the implementation of dematerialisation in particular, financial incentives for environmentally beneficial activities need to be promoted, while subsidies for unsustainable and environmentally problematic actions need to be abolished. Subsidies in an environmental context are defined as “all forms of explicit financial assistance to polluters or users of natural resources, e.g. grants, soft loans, tax breaks, accelerated depreciation, etc. for environmental protection” (OECD, 1999a: 9). As such, subsidies are a useful means to reduce market imperfections and to promote environmentally friendly technologies. On the other hand, in the form of ‘perverse subsidies’ they can exert “adverse effects on both the economy and the environment in the long run” (Myers/Kent, 1998: 8). A thorough assessment of their effectiveness is thus vital for the achievement of environmental objectives. The primary aim of policies dealing with subsidies should be to reduce those maintaining economic structures. Concerning sustainability, this would apply to all 15

those that prove ecologically harmful. Generally, all subsidies should be of a temporary nature, being reviewed at regular intervals (Hans-Böckler-Stiftung, 2000). All forms of subsidies have some impact on resource use, whether intended or not (Hinterberger et al., 1996). Those with apparent negative consequences include energy subsidies favouring unsustainable and environmentally problematic energy sources such as coal and nuclear energy, while penalising environmentally friendly sources such as biomass and other renewables. For example, “coal subsidies slowed down the shift to cleaner sources of energy production such as gas or wind farms because using coal remains artificially cheaper” (Gervais, 2002: 41). Tax concessions for logging, settlement, and ranching can have an accelerating effect on deforestation, species loss, as well as soil and water degradation. Also, pesticide subsidies advancing their excessive use can lead to human health problems as well as water pollution and an increase in the pesticide resistance of affected species, while subsidies for water resource development and water use can result in the overuse of water for industrial and other purposes (Schmidt-Bleek, 2001). The abolition of such subsidies has long been demanded, their existence repeatedly justified by reasons such as lower costs of production abroad, protecting domestic employment, securing domestic supply, etc. (Hinterberger et al., 1996). It is questionable whether these arguments are sufficient in view of the costs incurred by ‘perverse subsidies’ alone. As noted by Myers and Kent, they increase government expenditures resulting in higher taxes or larger budget deficits, diverting “government funds from better options for fiscal support” (Myers/Kent, 1998: 8). They also distort economies by undermining investment decisions and reducing “the pressure for businesses to become more efficient” (Ibid.), fostering “many other forms of environmental degradation, which apart from their intrinsic harm, act as a further drag on economies” (ibid.: 13). All these reasons give evidence of their economic and environmental violation against sustainability and of them counter-acting against the objective of dematerialising the economic metabolism by more efficient resource use. The reduction of ‘perverse subsidies’ is thus of primary importance, leading to a ‘double dividend’ in the form of an acceleration of sustainable development and an increase of government funds “available to give a new push to sustainable development (e.g. by investing in research and development or in public transport). Alongside the abolition of ‘perverse subsidies’, the introduction of new forms of subsidies dependent on the material intensity of an industry is suggested. Industries with low material input levels or industries striving for a reduction of material input levels would thus be eligible for subsidies, whereas industries with high material inputs would not. This would create new financial incentives to reduce resource use. The MAIA-Method for measuring the material intensity of products ‘from cradle to grave’, could serve as a general criterion for restructuring the subsidy system in such a direction (Hinterberger et al., 1996). However, due to the negative effects on economic efficiency associated with them, support measures of that kind should only be applied temporarily. An 16

introduction of broad and cost-intensive measures of support should be avoided, with environmental subsidies mainly focussing on cases of research and development, innovations and best practices. Other policy instruments to raise incentives towards more sustainable consumption and production patterns, such as taxes and certificates, are generally regarded to be superior over subsides, both in terms of economic efficiency and environmental effectiveness (Hans Böckler Foundation, 2000). 3.2

Ecological Fiscal Reform

Apart from restructuring the subsidies system, an ambitious environmental fiscal reform policy must include taxes and charges supporting resource efficiency. As defined by OECD (1999b) taxes in an environmental context refer to “any compulsory, unrequited payment to general government levied on tax bases deemed to be of particular environmental relevance” (OECD, 1999b: 56). EU legislation, on the other hand, does not offer a precise definition in the area of environmental taxation. In the approach taken by the Commission, it states that “one likely feature for a levy to be considered as environmental would be that the taxable base of the levy has a clear negative effect on the environment. However, a levy could also be regarded as environmental if it has a less clear, but nevertheless discernable positive environmental effect” (European Commission, 1997: 4). A levy in this context refers to both taxes and charges, even though a distinction is generally made between the two. This paper will mainly deal with environmental taxes, constituting the cornerstones of a comprehensive fiscal reform on a broad scale. The need for ecological taxes arises due to external costs not being accounted for by market prices. Their aim is to internalise those externalities by equating private and social costs of using nature for production, making “prices work for environment”. The EU has been committed to correct prices for energy consumption since the 2001 European Council in Gothenburg. Furthermore, the sixth EAP calls for the “promotion of sustainable production and consumption patterns”, particularly through “promoting and encouraging fiscal measures such as environmentally related taxes and incentives” (European Commission, 2001a: 5). Environmental taxes have thus been recognised as powerful tools to integrate environmental objectives into the economy, providing for a variety of benefits such as “economic incentives to reduce pollution and resource use” (Gervais, 2002: 42), “revenues that can be used for fiscal reforms” (Ibid.) and the stimulation of investment in the environment. Next to creating revenue that can be used to finance environmental activities or to reduce other taxes, environmental taxes create incentive effects via the price mechanism leading environmental goods to become more expensive, hence reducing demand and excessive consumption, as well as the associated pressure on natural resources. Additionally, considering that the environment (e.g. energy) may be regarded as another factor of production next to labour and capital, a rise in the price of the environment implies a decrease in the relative prices of the other factors. 17

A decrease in demand for the factor environment would thus be accompanied by an increase in demand for labour and capital. Environmental levies are usually preferable to regulation in terms of economic efficiency. Since they leave the choice to the economic agents to either reduce resource use or pay, they will most likely cause firms with low reduction costs to reduce resource use and therefore to avoid the tax, whereas firms that are faced with high reduction costs will tend to pay the tax. Environmental levies thus allow firms faced with the highest reduction costs to continue with resource intensive production practices, while encouraging less resource use over all. In both cases, firms will reduce environmental damage until equilibrium is reached where marginal abatement costs equal marginal costs of damage done to the environment, i.e. the tax level. Regulation, on the other hand, does not allow for that option to reduce resource use according to the internal cost structure of an enterprise. Furthermore, it only provides incentives for the reduction of resource use up to a prescribed standard, whereas taxes provide “a continuous stimulus for technological development, as any improvement has a continuing effect in lowering the tax burden” (Paleocrassas, 1999: 96) For the case of dematerialisation, the selection of energy (or CO2-emissions) and material input as tax bases seems particularly useful, representing good estimations of the potential for environmental harmfulness of the product under consideration. Even though previous suggestions highlighted the importance of energy taxes as the prime tool for ‘greening’ the tax system, it is now recognised that the concentration on fossil fuels alone will not be sufficient to considerably reduce the total material requirement of an economy. Hence the need for a complementary tax in the form of a material input tax (MIT). Energy Taxes One of the early concepts for an ecological tax reform, advanced by von Weizsäcker (von Weizsäcker et al., 1992), proposes a shift of the tax burden from labour to environmental load in general and energy sources in particular. Considering the overuse of natural resources and the underuse of human resources (i.e. high unemployment rates) in the EU, and the fact that 85% of taxation in the EU falls on human endeavour, such a shift “may contribute to increased employment alongside environmental improvements” (Gervais, 2002). This is referred to as a ‘double dividend’. An increase in energy prices would support a structural shift away from capital intensive, environmentally harmful industries to labour intensive industries with lower environmental impact (Wicke, 1993). It should be kept in mind, however, that the scale of that double dividend depends largely on the organisation of the labour market. For example, if the labour market is “very rigid on hire and fire regulations, the energy tax will have no employment effect until the labour market has been liberalised” (Paleocrassas, 1999: 100).

18

Energy taxes may thus serve as a strategic alternative for influencing energy use as well as the economic structure in ecologically positive ways, because energy intensive industries are largely characterised not only by higher emissions or risks, but also by higher transport volumes, larger quantities of domestic and industrial waste and increased water and land use (Jänicke/Mönch, 1990). Price signals to consumers associated with increased energy prices will support this development, providing consumers with the possibility to differentiate between energy intensive and less energy intensive products. Among the advantages of energy taxes over material input taxes is the fact that there are much less energy sources than matters to be considered as inputs. Energy taxes are thus believed to be easier to administer. Their introduction should, however, be accompanied by considerations on how to reduce other material flows. For example, when applied to energy sources, this would allow for a differentiation between material intensive and less material intensive ways to produce one kilowatt hour of electricity (Hinterberger et al., 1996). Material Input Tax (MIT) The MIT is a tax that uses material input as its tax base, placing a certain amount of money on each ton of total material requirement. Taxing all anthropogenically induced resource flows (except water and air1) needed to produce domestic goods and services as well as imported ones, leads to the internalisation of associated external costs and thus to higher input costs. As a consequence thereof, production will be faced with increased incentives to reduce material input (Hans Böckler Foundation, 2000). This will be done in the most efficient way as long as the costs of reducing material input are lower than the additional costs imposed by the MIT (Ibid.). Additionally, MIT acts as a tool for differentiating between resource intensive and less resource intensive products, since some of the associated costs will be shifted to consumers. The resulting price signals will have a positive effect on consumer choice for goods and services with a lower total material requirement, ceteris paribus. An alternative to taxing quantities of total material requirement has been brought forward by Reijnder, who argues that only virgin materials should be taxed. This would affect the input side as a whole, increasing specialisation in low loss recycling methods. The importance of such a tax is emphasised by the fact that “the focus on these low loss methods seems to have come to a standstill within the recycling industry” (te Riele et al, 2001: 20). Macroeconomic models for Germany have shown that dematerialisation by means of MIT is possible without frictions in the system (Hans Böckler Foundation, 2000). The results show that one of the conditions needed to fulfil the criteria of 1

Incentives for the reduction of water usage are to be posed by a progressive waste water charge, whereas the use of air is to be covered by the reduction incentives posed by a CO2 tax or its equivalent (Hans Böckler Foundation, 2000).

19

sustainable development – the need for increases in resource productivity to be higher than economic growth – can be satisfied by simulation. In one of these models, Panta Rhei, the economy has been divided into 58 production sectors, the total material requirement of which has been taxed (i.e. biotic and abiotic material inputs as well as soil erosion). The tax rate has been determined both according to the revenues generated as well as according to its effects on material flows. Instead of a sudden introduction of MIT at some finalised tax rate, a slowly and progressively rising MIT is recommended for economic agents to have enough time for adjustment. The announcement of future tax introductions/changes well in advance is of vital importance for the same reason. In a first step, MIT will affect only few industries that turn over large amounts of domestically mined or imported material, such as ore extraction, coal mining, building construction etc. Higher prices of material input and the associated incentives to reduce resource use will, however, precipitate through all parts of the product chain to consumers. As opposed to levies at the end of the life cycle of a product, MIT will thus be impossible to be evaded by enterprises, given European harmonisation (Hans Böckler Foundation, 2000). It should be noted, however, that a price increase of material input will only be effective where material input is an ecologically relevant tax base. It is clear that there are considerable differences among the damage coefficients of different materials. Therefore, a universal and undifferentiated MIT is deemed to lead to allocative inefficiencies and economically questionable structural effects (Hans Böckler Foundation, 2000). As such it is not entirely clear whether MIT is able to support an ecologically oriented structural change. In general, material intensive industries (especially in the primary sector) will be affected to the highest degree, whereas resource efficient services will be relatively advantaged. However, undifferentiated taxation of all material inputs also favours environmentally hazardous, yet ‘weight resource efficient’ technologies and products (Ibid.). Other open questions concerning the tax level, fiscal design in general, international trade and the introduction of complementary control measures (Ibid.), cause MIT to remain an economic instrument that needs a lot more research before being implemented on a broad scale. For the above reasons, no universal, undifferentiated MIT has been introduced in the EU so far. However, first steps have been taken by some Member States. Examples given by Bringezu (2002) include the Swedish Tax on Natural Gravel introduced in 1996; the Danish Raw Materials Act of 1997 (covering stone, gravel, sand, clay, lime, chalk, peat, top soil and similar deposits); and the introduction of the Aggregates Levy (on sand, gravel and rock) in the United Kingdom in April 2002. There is no doubt, that an ecological fiscal reform is a powerful tool to reduce material flows. It shows advantages in terms of efficiency and practicability, e.g. as compared to regulations or tradeable permits. However, a fiscal reform as outlined above will necessitate complementary reforms of the broader fiscal and financial system in order to be successful, including new approaches for depreciation rates, discounting to present value etc. (Paleocrassas, 1999). 20

3.3

Certificates Trading Systems

Environmental taxes, as laid out above, intervene in the market system by correcting the prices using environmental goods and services. They are not intended to create markets in which the quality of the environment can be traded. Generally referred to as certificates, the aim of permits, rights or quotas is rather to create marketable property rights for the use of environmental goods and services that would otherwise not be marketable. Public authorities can create markets for environmental ‘bads’ by limiting environmental harm of production (e.g. resource use, emissions etc.) to some agreed quantity and distributing that limit over tradeable certificates. Producers of environmental ‘bads’ would be required to own or purchase a quantity of certificates corresponding to their environmental use. Certificates are thus “based on the principle that any increase in emissions or in the use of natural resources must be offset by a decrease of an equivalent, or sometimes greater, quantity” (OECD, 1999b: 8). Even though they have so far mainly been implemented to reduce emissions, certificates have also been proposed in the form of material input (MI) certificates for limiting resource consumption. In such a system, MI-certificates would “constitute a permission to displace a certain quantity of primary material (following MAIA – in tons of MI)” (Hinterberger/Meyer-Stamer, 1997: 14). The design of that system could be established with reference to the two broad types of tradeable permit systems currently in operation for reducing emissions: one based on reduction credits, and the other based on ex-ante allocations (‘cap-and-trade’). The former approach “takes a ‘business as usual’ scenario as the starting point, and compares this baseline with actual performance” (OECD, 1999b: 8). Producers consuming fewer resources than anticipated by the baseline scenario would earn MI-credits which they could either use themselves for additional material input in the future or sell to other producers with higher resource consumption than accepted by the baseline. Of course, credits will only be bought when their price is lower than the required costs to reduce resource consumption (cp. OECD, 1999b). Since the price of the credits is solely determined by market supply and demand, credits will be most efficiently allocated to producers with the highest MI reduction costs. The ‘cap-and-trade’ approach “sets an overall emission/use limit (i.e. the ‘cap’) and requires all emitters to acquire a share in this total before they can emit [/use]” (OECD, 1999b: 8). The resource use limit would be introduced according to the macroeconomic reduction goal (e.g. a factor of X) and translated into permits that could either be allocated free-of-charge or auctioned by public authorities. Firms whishing to move primary material will then be required to “return a corresponding amount of certificates to the issuing authority in exchange” (Hinterberger/MeyerStamer, 1997: 14). However, in order to achieve a reduction in material input by a factor of ten within the next 50 years, the quantity of certificates would need to be 21

reduced by 5% each year (Hinterberger et al., 1996). The price increase associated with such a reduction in supply of certificates will result in further increased incentives to reduce material input, either by using less primary materials, more recycled materials or by switching to new product designs that require diminished resource flows (Ibid.). Certificates belong the group of – what Huppes/Simonis (2000) refer to as – market volume instruments. They focus on the allocation of a permitted quantity of resource use/emission while leaving its price up to the market. The price for the environmental ‘bads’ under consideration will thus be the result their supply and demand, i.e. where the marginal costs of reduction equals the marginal benefit of reduction. On the other hand, financial instruments (e.g. subsidies, taxes etc.) influence prices directly, leaving it up to the market to decide on the quantity of environmental ‘bads’ to be consumed at a given price. Again, at the right tax level marginal costs of reducing material input equal the benefits of that reduction. Whereas the latter are largely used on local, regional and national levels, market volume instruments play an increasing role in international environmental policy, especially climate policy. Of particular relevance to this paper are the four ‘Kyoto Mechanisms’ laid out in the Kyoto Protocol of 1997 for countries to achieve their targets for emission of greenhouse gases. Even though they act upon the output of the economic system, they can also serve as an impetus for the discussion on MIcertificates. The Kyoto Mechanisms consist of: •

•

•

•

‘bubbles’, where “a group of countries defines a joint target which is the sum of the original country targets and then redistributes the target among its members” (Michaelowa, 2001: 11); ‘Joint Implementation’, where “a country invests in emission reduction or sequestration projects in other countries with emission targets and thus earns ‘emission reduction units’” (Ibid.); the ‘Clean Development Mechanism’, a very elaborate instrument which allows for a country to earn ‘certified emission reductions’ “created through projects in countries without targets” (Ibid.); and ‘International Emissions Trading’, allowing for counties with emission targets to transfer ‘assigned amount units’ between each other (Michaelowa, 2001).

The bubble concept has been propagated to help existing firms/countries to achieve environmental targets in the most cost-effective way. They allow for producers to reduce the use of environmental goods in excess of the baseline reduction requirement in areas where the reduction is most cost-effective. The resulting surplus in reduction requirements can be balanced against higher environmental impacts of sources with reduction potentials that are more difficult and expensive to realise. Bubble policies thus allow for a greater flexibility in complying with market volume interventions leading to considerable cost savings (Wicke, 1993). 22

Cost reduction is also the aim of international trading schemes. Economic models have shown that Integrated Emissions Trading, as proposed in the Kyoto Protocol, can reduce costs by 30% to 90% for countries and companies off the cost of curbing emissions without trading (OECD/IEA, 2001). The saving potentials of an international MI-certificates trading scheme remain open, however, the market price for material input certificates will provide an incentive for the development of investment strategies and for seeking out appropriate technology to meet material input reduction targets at lowest possible costs (cp. OECD/IEA, 2001). Reducing resource use at low cost, given the fact that trading allows governments and businesses to reduce resource use wherever it is cheapest to do so (Ibid.) are some of the arguments brought forward by advocates of MI trading schemes. In comparison to taxes they additionally have the advantage of avoiding “the relatively arbitrary definition of a tax rate which may overshoot or underestimate external environmental costs” (Paleocrassas, 1999: 90). On the other hand, opponents of trading contend that such a system may generate unacceptably high levels of local resource use by allowing some sources to exceed their targets (cp. OECD/IEA, 2001). Furthermore, there are a number of questions that remain to be answered regarding “the definition, number, duration and spatial validity of the permits as well as the proposed method of their allocation” (Sterner, 2003: 83). Another problem arises due to the fact that economic growth and technological development may not always be in balance and the associated needs to adjust the system to an overall reduction in resource use, i.e. to reduce the supply for permits, allow market to collapse etc. (Paleocrassas, 1999). For the successful operation of a certificates trading system, it is of advantage to cover a large number of enterprises with similar resource use to facilitate the emergence of a market, independent of the size of the area. Furthermore, a sophisticated monitoring and measuring mechanism, both for industry and permit authority, is of vital importance, presumably leading to high monitoring and enforcement costs (Ibid.). However, the biggest challenge for an international trading scheme applying to all resources and all kinds of pollution will be “to define a world-wide basis variable per capita levels of sustainable consumption or tolerable pollution for each resource, each pollutant and each region” (Paleocrassas, 1999: 90) and the following task of “negotiating the appointment of property rights among local, regional, national and world population” (Ibid.). It remains open whether these obstacles can be overcome, and how far the above instruments can be integrated into a MI-certificates trading scheme, however, a cooperation of the international community seems sensible in view of the increasing interaction of national economic systems and resulting interlinkages of material flows.

23

3.4

Public Procurement

An often neglected instrument to reduce the domestic material requirement is public procurement. According to Gervais (2002), public purchases of services, works and goods within the EU account to over € 1,000 billion annually or about 14% of the EU’s GDP. Such figures “confirm the importance of public procurement, in terms of impact on the Single Market and competition, as well as opening of markets for European suppliers” (EU, 2003). They also emphasise the leading role such a large sector can play to “boost innovation and development and [to] encourage the offer/demand for environmentally friendly products and services” (Gervais, 2002: 45). This also applies for eco-efficient products with low material input per service unit. However, EU public procurement legislation guaranteeing fair access for suppliers sets limits on governments for attaching environmental conditions to their purchasing contracts. Furthermore, even within the limited scope to benefit the environment, “any such criteria, at this stage, must provide benefits to the contracting authority rather than to the wider community” (Ibid.). This also applies for environmental externalities, which by definition are born by society as a whole. These costs could only be taken into account in specific cases, “for instance where external costs are due to the execution of the contract and at the same time are born directly by the purchaser of the product or service in question” (European Commission, 2001b: 22) On the other hand, by stipulating precise green requirements in the contract specification, authorities “could take into account the ‘environmental soundness’ of products or services, for example, the consumption of natural resources, by ‘translating’ this environmental objective into specific, product-related and economically measurable criteria by requiring a rate of energy consumption” (European Commission, 2001b: 20). For the assessment of the most economically advantageous tender, the life-time costs of a product to be born by the contracting authority may be considered. Running costs and cost effectiveness as possible award criteria “might include direct running costs (energy, water and other resources used during the lifetime of the product); spending to save (for example, investing in higher levels of insulation to save energy and thus money in the future); as well as the costs of maintenance or recycling of the product” (European Commission, 2001b: 21). To promote the case of dematerialisation in public procurement, the legal framework allowing for governments to favour eco-efficient products and services where possible should be utilised to its full potential.

4.

Regulatory Instruments

Environmental policy has traditionally been dominated by regulations, and continues to be so even after the introduction of market oriented and voluntary instruments towards the end of the 20th century. Regulatory instruments are defined as “a 24

category of environmental policy whereby public authorities mandate the environmental performance to be achieved or the technologies to be used” (EEA, 2004). They are distinct from law, insofar as regulations may be considered as law put into practice by public authorities. Regulations may thus be regarded as rules “that are designed to fill in the details of the broad concepts mandated by the legislature in statutes” (New York State, 2004). As such, they seem to be predestined to serve the goal of dematerialisation. However, as will be shown, this only applies to specific situations. Belonging to the group of budget neutral measures, regulatory instruments are often chosen over economic ones “where there is little flexibility allowable on the timing or nature of the outcome required” (DEFRA, 2002: 2). This applies primarily to “situations where high environmental risk and irreversibilities of environmental impacts call for rapid preventive action” (Bartelmus et al., 2000: 37). Health concerns and ecotoxicity are practical examples of cases where “regulation has to be binding and enforceable throughout a country and its population according to the same standards” (Spangenberg et al., 1999: 4). Often referred to as command-and-control measures, regulations are associated with specific targets for the reduction of environmental use, and the threat of sanctions in cases of non-compliance. Violations of regulations may be sanctioned with civil or criminal penalties or both, depending on the mandate given by law (New York State, 2004). Sterner (2003) distinguishes between three kinds of direct regulation of the environment: the direct provision of public goods, regulation of technology, and regulation of performance. Even though not always recognised as a proper regulatory instrument, the direct provision of public goods may be considered as the most straightforward policy to solve specific environmental problems. It refers to the application of personnel, know-how, and resources of a public authority towards improving the environment. Examples include taking responsibility for major environmental threats, managing certain kinds of research and control functions etc. (Sterner, 2003). However, due to recent efforts to privatise government enterprises, the role of the state has started to decompose from the direct production of goods and services to individual components, namely financing, administration, provision, and control. Due to the fact that some or all of these components may be transferred to the private sector, the direct provision of public goods represents only one end of the spectrum of organising public goods (Ibid.). The regulation of technology, on the other hand, represents “one way of regulating the behaviour of firms, households, agencies, and other agents in the economy by prescribing the technology to be used or the conditions” (Sterner, 2003: 75). The prescribed technology is also referred to as best practical technology, best available technology etc. Next to prescribing technology, technology regulation may also be used to restrict timing or location of the use of certain technologies. In that sense, a ban forbids specific processes or products that are deemed to be of special 25

importance for resource use. Similarly, zoning refers to certain methods or technologies being banned or limited to a certain area (Sterner, 2003). Even though the simplicity of technology regulations needs to be recognised together with their short term effectiveness, they lack the flexibility for enterprises to choose the most cost-effective way to reduce resource use. Additionally, mandatory technology alone will not be sufficient to reduce resource use any further than allowed by best available technology (Ibid.). It also does not recognise the need to provide incentives for the development of new technology. Technology regulations may, however, be the instrument of choice where clarity of decision-making meets economy of control and ease of monitoring. A preference for resource efficient technologies can serve as an important factor for reducing material input. For example, producers may be obliged – where possible – to supply industrial waste heat to district heating networks, thus reducing the pressure on primary materials (Wicke, 1993). Instead of requiring a particular technology, performance standards can be used to limit harvests or emissions to certain quantities. Analogously, they may also be applied to restrict the use of material input, leaving it up to the producer to select the most appropriate (i.e. cost-effective) means to reach that target. Input regulations on the production process can oblige producers to select only a limited amount of natural resources for their activities. As such, they may focus on quantities to be limited or reduced as has been the case in “fleet efficiency regulations or licences for mining (relative input limitations) and logging or ground water extraction permits (usually absolute input limitations)” (Spangenberg et al., 1999: 3). They may also prohibit the use of certain resources that are associated with a high environmental burden, focussing on qualities rather than on quantities. Additionally, regulations can impose production quotas on economic agents, which can go as far as prohibiting production at all. Such policies are mainly used with regard to emissions, but may also be useful for highly material intensive products. Even though still widely implemented for output related environmental protection, regulatory instruments are losing importance due to lack of flexibility and incentives to reduce environmental use to a level lower than imposed by public authorities. Furthermore, they often do not take into account individual marginal abatement or reduction costs, leading to an economically inefficient attainment of environmental targets at unnecessarily high costs. However, the efficiency of regulatory systems can be considerably increased by complementing them with economic incentives (Costanza et al., 1997). Due to the above deficiencies and the fact that de-regulation is ranking high on the political agenda (Pearce, 2000), it does not seem appropriate to recommend regulatory instruments on a broad basis. However, environmental legislation needs to continue to play an important role in environmental policy since “you cannot stop treating environmental damage that endangers human health as a criminal offence. You can make it a civil offence too, giving rise to compensation to the damaged 26

party, but it must always remain basically a criminal offence” (Paleocrassas, 1999: 97). Additionally, regulations are necessary to provide for a general regulatory framework which is needed for markets to function efficiently, dealing with institutions, public participation, transparency, technology transfer etc. as laid out by the WSSD in Johannesburg. As such they may specifically serve the promotion of corporate responsibility and accountability. With regard to resource use, regulations have been recommended to focus on resource intensive industries and intermediate products (Schütz/Welfens, 2000). Together with other quantity based approaches (e.g. tradeable quotas) they have also been recommended for the promotion of strong sustainability, due to the higher certainty of achieving dematerialisation targets as compared to price based approaches, such as taxes or charges (Pearce, 2000). In general, regulations are best suited in situations “where there are clear environmental goals with overwhelming political consensus, similar costs of abatement across all actors, relative certainty about what is being emitted [or about material input and associated rucksacks], and easy and effective enforcement” (Costanza et al., 1997: 196). The need to adjust the regulatory framework for dematerialisation is evident. This also applies to supplementing and/or replacing existing output related regulations “by the development of input policies […] in order to reduce the total throughput of our economies” (Spangenberg et al., 1999: 4).

5.

Recommendation for a Policy-Mix

Many policy instruments suited to bring about a reduction of material flows in the EU and elsewhere have been presented in this chapter. The range has covered voluntary instruments, with little government intervention; economic instruments aimed at increasing incentives to reduce material flows; and regulatory measures marked by high levels of government intervention. In order to reach sustainability and associated material flow targets a single instrument operating in isolation will usually not be enough. Rather, a package of instruments covering a wide selection needs to be implemented, allowing for different problems to be combated by different solutions. This requires diverse types of instruments to work alongside: new instruments like certificates together with old ones like subsidies, with some of them having an effect in the long run, others in the short run. It may also be required for such a policy-mix to change over time. A package of instruments could, for example, “require as a minimum specific changes in behaviour; encourage greater changes in behaviour in the short run; incentivise even greater changes in behaviour in the long run; and facilitate those changes, at least in the short run” (DEFRA, 2002: 4). Environmental policy instruments need thus to be analysed as to the time horizon of their impact, in order for a policy mix to be effective in the short run as well as in the long run. Within this context, the short run is determined by the possibility for technological, environmental and institutional changes with given technology, 27

urban structure and institutional set up. In comparison, the long run requires structural changes within these elements for additional developments to take place (Paleocrassas, 1999). Regarding resource taxes, the impact of consumption resource taxes is more limited than that of purchase or investment resource taxes, especially in the short-run (Ibid.). This results from the fact that producers generally “have higher price elasticities than consumers because they are more cost conscious” (Paleocrassas, 1999: 98). It thus follows that production patterns are easier changed than consumption patterns, legitimising the use of energy taxes and material input taxes to achieve dematerialisation within shorter time horizons. This will lead producers to exhaust all technological possibilities to reduce energy and resource use within the short run, while providing a continuous stimulus to technological development in the long run. Tradeable permits, on the other hand, will only have an impact on resource use in the long run, due to long design and gestation periods and functioning markets needing time to gain depth and experience (Paleocrassas, 1999). This is especially relevant with regard to the proposed introduction of a comprehensive international system covering all resources, which would need an extensive period of preliminary negotiations between members of the international community. Regulatory instruments usually have an immediate effect, starting from the date of their application, even though they are often provided with some delay mechanisms that reduce the force of their impact (Ibid.). In these cases the date of adoption deviates from the date at which the regulations come into force. As such, regulations may be useful in the short run, however, in order to provide periods of adjustment for producers, they should be introduced with a long run orientation. This orientation also requires anticipated revision dates or, where appropriate, expiration dates. Regulatory instruments are best suited to keep the environmental impacts of resource use within the general ‘guard-rails’ of sustainability in the short and long term by targeting on the reduction of environmental risks and irreversibilities of environmental impacts associated with specific substances. For the unspecific reduction of material flows, as promoted by the concept of dematerialisation, a policy mix emphasising economic instruments such as material input taxes is suggested (Hans Böckler Foundation, 2000). They are designed to promote a shift away from material intensive activities, providing for long term incentives for reducing the use of primary materials. This shift may also be facilitated though the availability of funds generated by taxes (cp. DEFRA, 2002). A mix of instruments is also recommended from the point of view that different desired ranges of environmental quality or ecological health require separate appropriate policy levels. This transdisciplinary approach, brought forward by the aforementioned field of ecological economics, “supplements economic insights through a team approach by explicitly including concepts from ecology and the physical sciences as well as concerns for equity, distribution and political feasibility” 28

(Costanza et al., 1997: 218). It thus goes beyond focussing on marginal cost functions by suggesting instruments with stronger government intervention in situations where the consumption of environmental goods and services exceeds the limits of sustainable development. The ecological economics approach to pollution control, as illustrated in Figure 1, can serve as an important impetus for the discussion about reducing material flows. Figure 1: An Ecological Economic Approach to Pollution Control

Source: Costanza et al., 1997 Following this approach, regulatory instruments are best suited for producers whose resource use exceeds ecologically acceptable limits and where material input (and associated output) leads to irreversible, nonsustainable damage to the system (cp. Costanza et al., 1997). They thus prohibit unconstrained resource use, serving “as a safeguard against miscalculation and uncertainty” (Costanza et al., 1997: 220). On the other end of the spectrum lies the property rights zone, where material flows are too low to be measured or whose effects on the productivity of the system are negligible. Within this range, it is expected that marginal costs of monitoring and administration of environmental policy would exceed the marginal benefits of reducing material input, thus not justifying government intervention (cp. Costanza et al., 1997). Producers may therefore be allowed to use resources within legal limits without being charged for their behaviour. For reasons of equity and efficiency, this could also be implemented by means of tradeable permits, which would be offered for free within the property rights zone, and offered for sale thereafter.

29

Between these two zones lies the incentive zone, associated with measurable damages to the environment and the productivity of the system. For this zone, the implementation of economic instruments such as charges or taxes is recommended, providing financial incentives through market forces for producers to reduce resource use. Producers would be endorsed to “shift entrepreneurial talents away from regulatory evasion toward efficient, less entropic technical improvement” (Costanza et al., 1997: 219) as they progress from the regulatory zone to the incentive zone. Even though in practice not as straight forward as it seems in theory, the ecological economics approach to implementing environmental policy instruments is useful for approximating relevant instruments for particular situations, diverging from “the strict efficiency rule of taxing each unit of emission [or material input] at the same price in order to provide some equity consideration to emitters” (Costanza et al., 1997: 222). Independent of the actual design of the policy mix, it needs to be complemented by a sophisticated system of environmental monitoring and control. This does not only apply to regulatory instruments but especially for the introduction of new market-based instruments. Associated with an increase in the efficiency of economic instruments, environmental monitoring and control allow for the reduction of enforcement deficits, as well as for enterprises to adjust to new environmental requirements in due time. It may thus be in the interest of producers to monitor their own enterprises, not only – where possible – to escape public controlling measures, but also to obtain an overview of their own material requirement. The latter can reduce the regulatory pressure on producers, allowing for additional time in preparing for interventions of public authorities, which may be used for searching for the most cost-effective solutions (Wicke, 1993). In conclusion, taking into account the complexity of environmental problems it does not seem sensible to concentrate on particular environmental policy instruments for achieving the goal of dematerialisation. A suitable blend of instruments needs thus to be developed, taking care of the general problem of unsustainable material flows, as well as the dangers associated with the use of specific resources. It should be the result of a transparent political process taking into account ecological, as well as economic and social objectives in order to ensure the sustainability of Europe’s development in the future.

30

References AIM – Association des Industries de Marque (1997) Economic Instruments in Environmental Policy, AIM Briefing Paper, September 1997 Altmann, J. (1997) Umweltpolitik: Daten, Fakten, Konzepte für die Praxis (Stuttgart, Lucius und Lucius) Bartelmus, P./Bringezu, S./Moll, S. (2000) Dematerialization, Environmental Accounting and Resource Management – main issues and how they can be translated in to public policy initiatives, Expert Report on Resource Management commissioned by the European Commission, Directorate General Environment Begg, D./Fischer, S./Dornbusch, R. (1994) Economics – Fourth Edition (Berkshire, McGrawHill) Bringezu, S. (2002) Towards Sustainable Resource Management in the European Union, Wuppertal Papers No. 121 (Wuppertal, Wuppertal Institute for Climate, Environment, Energy) Costanza, R./Cumberland, J./Daly, H./Goodland, R./Norgaard, R. (1997) An Introduction to Ecological Economics (Boca Raton, St. Lucie Press) Croci, E./Pesaro, J. (1999) Voluntary Agreements in the Environmental Sector – The Italian Experience, CAVA Working Paper No. 98/11/5 (European Research Network on Voluntary Approaches for Environmental Protection) DEFRA (2002) The Government’s Strategic Review of diffuse water pollution from agriculture in England: Types of Environmental Policy Instruments (London, Department for Environment, Food and Rural Affairs) EEA (2002) Environmental signals 2002 – Benchmarking the millennium, European Environment Agency regular indicator report (Luxembourg, Office for Official Publications of the European Communities) EEC (1993) ‘Council Regulation (EEC) No 1836/93 of 29 June 1993 allowing voluntary participation in the industrial sector in a Community eco-management and audit scheme’ in Official Journal of the European Communities, No. L168, Volume 1 EPA (1998a) Extended Product Responsibility – A Strategic Framework for Sustainable Products (Washington D.C., EPA) EPA (1998b) ‘Extended Product Responsibility’ in WasteWise Update, October 1998 (Washington D.C., EPA) European Commission (1997) Environmental Taxes and Charges in the Single Market, Communication from the Commission, 26 March 1997 (Brussels, European Commission) European Commission (2001a) On the sixth environmental action programme of the European Community - Environment 2010 - Our future, our choice, COM(2001) 31 final, 24 January 2001 European Commission (2001b) Commission Interpretative Communication on the Community law applicable to public procurement and the possibilities for integrating environmental considerations into public procurement, COM(2001) 273 final, 4 July 2001

31

European Commission (2002) Communication from the Commission to the European Parliament, the Council, the Economic and Social Committee and the Committee of the Regions – Environmental Agreements at Community Level Within the Framework of the Action Plan on the Simplification and Improvement of the Regulatory Environment, COM(2002) 412 final, 17 July 2002 Gervais, C. (2002) An Overview of European Waste and Resource Management Policy (London, Forum for the Future) Hans Böckler Stiftung (2000) Verbundprojekt Arbeit und Ökologie Abschlussbericht (Düsseldorf, Der Setzkasten GmbH) Hinterberger, F./Luks, F./Stewen M. (1996) Ökologische Wirtschaftspolitik – Zwischen Ökodiktatur und Umweltkatastrophe (Berlin, Birkhäuser Verlag) Hinterberger, F./Meyer-Stamer, J. (1997) Knowledge and Environment – Innovation for Sustainable Development, CompETE Working Paper No. 2, December 1997 (Competitiveness, Employment, Technology and Environment) Huber, R. M./Ruitenbeek, J./da Motta, R. S. (1998) Market Based Instruments for Environmental Policymaking in Latin America and the Caribbean – Lessons from Eleven Countries, World Bank Discussion Paper No. 381 (Washington D.C., World Bank) Huppes, G./Simonis, U. E. (2000) Environmental Policy Instruments in a new era (Leiden, Leiden University) James, D. (1997) Environmental Incentives, Australian Experience with Economic Instruments for Environmental Management, Environmental Economics Research Paper No. 5 (Canberra, Environment Australia) Jänicke, M./Kunig, P./Stitzel, M. (1999) Umweltpolitik (Bonn, Dietz) Jänicke, M./Mönch, H. (1990) Ökologische Dimensionen wirtschaftlichen Strukturwandels, Zwischenbericht, FFU-Report 90-10 (Berlin, Freie Universität Berlin) Michaelowa, A. (2001) Rio, Kyoto, Marrakesh – Groundrules for the Global Climate Policy Regime, HWWA Discussion Paper No. 152 (Hamburg, Hamburg Institute of International Economics) Moll, S./Bringezu, S./Schütz, H. (2003) Resource Use in European Countries (Copenhagen, European Topic Centre on Waste and Material Flows) Myers, N./Kent J. (1998) Perverse Subsidies – Tax $s Undercutting Our Economies and Environments Alike (Winnipeg, IISD) OECD (1999a) Voluntary Approaches for Environmental Policy – An Assessment (Paris, OECD) OECD (1999b) Economic Instruments for Pollution Control and Natural Resources Management in OECD Countries: A Survey (Paris, OECD) OECD (2001a) Extended Producer Responsibility – A Guidance Manual for Governments (Paris, OECD) OECD (2002) Indicators to Measure Decoupling of Environmental Pressure from Economic Growth (Paris, OECD) OECD (2003) Voluntary Approaches for Environmental Policy (Paris, OECD) OECD/IEA (2001) International Emission Trading – From Concept to Reality (Paris, OECD/IEA) Ökobüro/ÖGUT (2003) Environmental Governance und Umweltvereinbarungen (Vienna, Ökobüro/Österreichische Gesellschaft für Umwelt und Technik)

32

Paleocrassas (1999) Fiscal Reform, Resource Productivity, And Employment, Factor 10 Report, Chapter II (Factor Ten Institute) Pearce, D. (2000) Public Policy and Natural Resources Management – A Framework for Integrating Concepts and Methodologies for Policy Evaluation, September 2000 (London, University College London) Schmidt-Bleek (1996) MAIA – Einführung in die Materialintensitäts-Analyse nach dem MIPSKonzept (Wuppertal, Wuppertal Institute for Climate, Environment, Energy) Schütz, H./Welfens, M. (2000) Sustainable Development by Dematerialization in Production and Consumption – Strategy for the new Environmental Policy in Poland, Wuppertal Papers No. 103 (Wuppertal, Wuppertal Institute for Climate, Environment, Energy) Spangenberg, J. H./Hinterberger, F./Moll, S./Schütz, H. (1999) Material Flow Analysis, TMR and the mips-Concept: A Contribution to the Development of Indicators for Measuring Changes in Consumption and Production Patterns (Wuppertal, Wuppertal Institute for Climate, Environment, Energy) Sterner, T. (2003) Policy Instruments for Environmental and Natural Resource Management (Washington DC, RFF Press) Stewen, M. (2000) Inputorientierte Umweltpolitik – Von der Idee zur Umsetzung, Wiener Arbeitspapier zur ökologischen Wirtschaftspolitik Nr. 3 (Vienna, Bureau for Ecological Economic Policy) Stewen, M. (2002) Inputorientierte Umweltpolitik in der Sozialen Marktwirtschaft (BadenBaden, Nomos Verlagsgesellschaft) Te Riele, H./van Elburg, M./Kemna, R (2001) Dematerialisation – Less clear than it seems (The Hague, Dutch Ministry of VROM) UNEP (2002a) Global Environmental Outlook – 3 (London, Earthscan Publications Ltd) UNEP (2002b) Industry as a Partner for Sustainable Development – 10 Years After Rio: the UNEP Assessment (Paris, UNEP) UNESCO-UNEP (1978) ‘The Tbilisi Declaration’ in Connect, UNESCO-UNEP Environmental Education Newsletter, Vol. III, No. 1, January 1978 Von Weizsäcker, E. U./Jesinghaus, J./Mauch, S./Iten, R. (1992) Ökologische Steuerreform. Europäische Ebene und Fallbeispiel Schweiz (Chur/Zurich, Verlag Rüegger) WCED (1987) Our Common Future (Oxford, Oxford University Press) Wicke, L. (1993) Umweltökonomie (Munich, Verlag Vahlen) Electronic Sources It should be noted, that electronic sources are quoted without page references. Please refer to the corresponding web pages for further information. Carter, N. (2001) ‘Understanding Sustainable Development’ at http://www.fathom.com/course/21701763/session2.html (as of 12 January 2004) EEA (2004) ‘Regulatory Instrument’ at http://glossary.eea.eu.int/EEAGlossary/R/regulatory_instrument (as of 7 January 2004) EPA – Environmental Protection Agency (2003) ‘Environmental Education Improves Our Everyday Lives’ at http://www.epa.gov/enviroed/ (as of 9 September 2003) EU (2003) ‘Public procurement in the European Union’ at http://europa.eu.int/business/en/topics/publicproc/ (as of 1 October 2003)

33

IISD – International Institute for Sustainable Development (2003) ‘The Evolution of Economic Instruments’ at http://iisd1.iisd.ca/greenbud/evolutio.htm (as of 6 October 2003) Nachhaltigkeit (2001) ‘Das Monatsthema: Dematerialisierung, Ökoeffizienz und Zero Emission’ at http://www.nachhaltigkeit.at/bibliothek/pdf/monatsthemen2001/mainM01_05.pdf (as of 12 January 2004) New York State (2004) ‘What are Regulations’ at http://www.dec.state.ny.us/website/dshm/regs/whyregs.htm (as of 7 January 2004) UNEP (1994) The Use and Application of Economic Instruments for Environmental Management and Sustainable Development, Environmental Economics Series Paper No. 12 at http://www.unep.org/unep/products/eeu/ecoserie/ecos12/ecos121.htm (as of 21 September 2003)

34