Selection and design of environmental policy instruments✶

ARTICLE IN PRESS

Selection and design of environmental policy instruments✶ ∗ Department

Thomas Sterner∗,1 , Elizabeth J.Z. Robinson†

of Economics, University of Gothenburg, Gothenburg, Sweden of Agriculture, Policy, and Development, University of Reading, UK 1 Corresponding author: e-mail address: [email protected]

† School

CONTENTS 1 The Need for Policy ............................................................................. 2 Policy Failures ................................................................................... 3 The Menu of Instruments ....................................................................... 3.1 Price-Type Instruments ........................................................... 3.2 Rights-Based Policies ............................................................. 3.3 Regulation .......................................................................... 3.4 Information or Legal-Based Policies ............................................ 3.5 The Process of Policy Making at National or Other Levels .................. 4 The Selection of Instruments .................................................................. 4.1 Efficiency ........................................................................... 4.2 Information Asymmetries and Uncertainty..................................... 4.3 Intertemporal Efficiency .......................................................... 4.4 Spatial Efficiency .................................................................. 4.5 Practical and Political Aspects .................................................. 4.6 Normative Principles, Distributional Aspects, and Environmental Justice 5 Selected Examples .............................................................................. 5.1 Taxing Carbon ...................................................................... 5.1.1 Effects of CO2 Taxation ......................................................... 5.2 Taxing (and Subsidizing) Transport Fuel ....................................... 5.3 Cap and Trade Schemes .......................................................... 5.4 Refunding Emission Payments .................................................. 5.5 Regulation Versus Taxation: The Example of a Hazardous Chemical....... 5.6 Policies to Modify Behavioral Norms ...........................................

2 4 6 7 9 10 11 12 13 14 15 17 18 19 21 23 24 26 28 30 32 34 37

✶ We would like to thank the editors and two very knowledgeable and generous referees for good comments. We would also like to thank Kristin Seyboth for exceptional editorial assistance. Financial support from Mistra Carbon Exit and the University of Gothenburg (UGOT) Centre for Collective Action Research (CeCAR) is gratefully acknowledged. Handbook of Environmental Economics, ISSN 1574-0099, https://doi.org/10.1016/bs.hesenv.2018.08.002 Copyright © 2018 Elsevier B.V. All rights reserved.

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6 Designing Policies for the Anthropocene.................................................... 6.1 An Expansion of Geographic and Political Scope ............................. 6.2 Significant Extension in Time-Scale ............................................ 6.3 Significant Extension of the Number of Pollutants and Scientific Complexity .......................................................................... 6.4 Equity, Ethics, Risk, Uncertainty, and Governance ........................... References............................................................................................

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1 THE NEED FOR POLICY In economics, we often assume each consumer or firm chooses actions that are optimal for that actor. If all non-market interactions are assumed away, the first welfare theorem states that the outcome is an efficient allocation of resources. Government policy cannot improve upon this ‘perfect’ market system, (only on distribution of assets) and may in fact be damaging. In such a ’perfect’ world, there would be no environmental problems – suggesting a potential libertarian bias. Reality, however, is not well described by a perfect market populated by rational “homines economici.” Rather, non-convexities are pervasive. These include externalities, public goods, poorly defined property rights, noncompetitive markets, increasing returns to scale, and asymmetric information, all of which are compounded by uncertainty. These market failures are conditions under which the free market fails to reach an optimal outcome. We need to understand the nature and extent of such imperfections in order to suggest remedies in the form of policy instruments. Externalities are side effects of production or consumption. The unregulated market oversupplies negative externalities, such as smoke or polluted water that can have significant yet avoidable health and welfare consequences, and undersupplies positive externalities. Cropper and Oates’ (1992, p. 678) early review of environmental economics suggests that “the source of the basic economic principles of environmental policy is to be found in the theory of externalities”. Public goods are products or services, including “ecosystem services”, enjoyed in common, such as a healthy atmosphere or the pollination services of bees, which are undersupplied by the market. Similarly, world knowledge is a global public good, thus investment in cleaner energy R&D may be undersupplied due to international knowledge spillovers (Bosetti et al., 2008). Common pool resources are those with high subtractibility, for which it is costly to exclude people, and where individual and group interests may differ (Buck, 2017). In these cases, ownership and management may be best undertaken collectively, such as by a group of people in a village or local community, as eloquently described by Ostrom (1990). Poorly defined property rights also create imperfect conditions. If rights are not well-defined for all possible goods and services, externalities result – as with neighbors who create disturbances, blocking a view or access to water, or producing unpleasant odors or noise. If all these aspects were governed by well-defined property rights (and there were low or zero transaction costs), then in principle external effects

ARTICLE IN PRESS 1 The Need for Policy

would be avoided (Coase, 1960). Yet it would be impractically costly and probably impossible to define and enforce sufficiently detailed property rights to cover every aspect that might give rise to external effects. Noncompetitive markets are those in which individual actors such as international companies are large enough to have market power and therefore undue influence on prices and rules in the marketplace. If there are increasing returns to scale or indivisibilities, implying production sets are not smooth or convex, then markets without government intervention are likely to fail to maximize social welfare. Extreme cases are monopolies, whether resulting from scale economies or protection. Asymmetric information, the lack of equal access to information, and incomplete information can each stop the market from operating perfectly (see Ulph, 1998, 2000, for examples of asymmetric information and federal systems). An example of asymmetric information is in the design of contracts for payments for ecosystem services, where land owners typically know more about the costs of contractual compliance than do the buyers of the conservation service (Ferraro, 2008). An example of where incomplete information leads to market failure is with respect to investment in innovations linked to climate change, where uncertainty is particularly acute (Jaffe et al., 2005). According to the theory of second-best, easy inferences can be drawn if there is just one (small) market failure. In reality, however, market failures can be plentiful and large, in which case it is very complex to know whether correcting one market failure at a time will move us progressively in the right direction. Market failures are one key motivation for environmental policy, but there are others that are also relevant. For example, behavioral biases and anomalies. Behavioral economics helps to explain why individuals do not always simply maximize utility and why individual behavior may be more complex than can be described by economists’ simple models, ultimately making it harder to evaluate welfare implications. When people act in groups, they may feel envy, altruism, anger, revenge, and many other feelings that can influence behavior. These insights help us understand why, for instance, some communities are able to overcome incentives to free-ride on the efforts of others and instead collaborate to manage a local resource, when standard economic theory of utility maximization would suggest the opposite (see for example, Cárdenas, 2000, 2016). This also encourages economists to consider aspects such as non-standard incentives, collective action, network externalities, and social norm externalities. Many market failures can be described as a so-called ‘prisoner’s dilemma’. The natural outcome of selfish rational decision making is that both individuals are worse off than if they trusted each other to behave in the way that makes the group better off as a whole. The theory of repeated games demonstrates that, with a sufficiently small discount rate, it may be individually rational to “cooperate”, such that all are better off, rather than “cheat”, which yields short-term individual gains but long-term losses to all. Yet game theory teaches us little about how to reach the cooperative equilibrium. Experimental economics and societal observations can give us some understanding of how society can reach a “better” equilibrium through changing in-

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dividual and group norms (Nyborg et al., 2016). This builds on the recognition that individual wellbeing depends on individual behavior relative to societal norms, and that societal norms can encourage individuals to make choices that either benefit or harm society and the environment as a whole. These critiques of neoclassical economics (with its emphasis on individually “rational” behavior, convex preferences, and a world in which absolutes rather than relatives matter) are deep and damning. There is increasing awareness that neoclassical assumptions – such as the concept of the “perfect market” and the efficiency of the “invisible hand” – are the exceptions rather than the norm. As a result, economists are increasingly recognizing and incorporating concepts of happiness, relative consumption, and utility functions that include the wellbeing of others (examples linked to environmental considerations include Welsch, 2006; and Ferrer-i-Carbonell and Gowdy, 2007). If, for example, social comparison and adaptation based on past experiences matter, then optimal environmental quality is greater than standard neoclassical economics would predict, implying that stricter environmental regulation is needed where relative consumption matters for environmental policy (Welsch, 2009). Some claim that the market fails to manage ecosystem resources because of vagaries of human “nature” such as ignorance, greed, and procrastination (see e.g. Naess, 1989). These are powerful forces indeed, but the main reasons why we have environmental disasters and overuse our resources are market (and policy) failures (see Section 2), as in collective action dilemmas where the incentives for individual action fail to align with societal interests (Ostrom, 1998). The picture of the perfect market as a safe norm and market failures as small exceptions needs to be abandoned. The perfect market is in fact, the strange abstraction (see Nyborg, 2016, for an entertaining illustration). Coordination mechanisms such as social norms may spontaneously evolve and do sometimes partially address the problems (Ostrom, 2000). Public policy can strengthen or create new norms and these can be seen as policy tools that help tip us toward a good equilibrium, for instance by internalizing potential positive network externalities or helping overcome problems of coordination and monitoring in common pool resource use (see further, Nyborg, 2016, 2017). The abstract case for policy instruments to correct for market failures is clear, but often their scale and complexity is hard to grasp and not easy to analyze with the tools of economics alone. There is a need to bring together the natural and physical sciences with economics to enable policy appraisals to take explicit account of the spatial variability and dynamics of the natural environment, and thus the feasibility and cost of implementing environmental policy.

2 POLICY FAILURES Where markets fail, policies should be able to rectify or at least dampen the extent of the failure. Yet many times these market failures are instead exacerbated by bad pol-

ARTICLE IN PRESS 2 Policy Failures

icy. Policy failure is not unique to environmental policy and there is a broad literature in political science concerning policy failure in general; see, for instance, McConnell (2015) or Derwort (2016). Fisheries provide many such examples. The absence of property rights leads to a tragedy of open access, and the situation is often compounded by policy failure. When there is overfishing and fishers are impoverished because of small catches, public policy should reduce fishing effort. Instead, politicians typically hand out subsidies intended to “help” fishers. These handouts may benefit fishers in the short run, but typically exacerbate the depletion of fish stocks, resulting in even higher costs and lower earnings. Bad policies thereby exacerbate the market failure, because those subsidies are typically spent on bigger boats and longer nets that speed up stock depletion. For examples, see, for instance Sanchirico and Wilen (2007), Khalilian et al. (2010), and Costello et al. (2016). Conversely, policy failures may also be a consequence of poor governance systems. The State, or the Government, is normally the ultimate designer and enforcer of institutional arrangements, including environmental policies. Both “input” and “output” characteristics of a state-centered political system potentially drive predatory policies and policy failures (Sjöstedt and Jagers, 2014). Input characteristics are defined as the procedures preceding a policy decision, such as the type and level of democracy and citizens’ access to public authorities. Output characteristics refer to the way in which that authority is exercised. On the one hand, lack of highstandard input components increases the risk of both weak policies and eventually poor compliance. On the other hand, low levels of impartiality in the exercise of public authority can generate policy failure, because factors such as corruption and poor government effectiveness negatively affect the public’s compliance with laws and regulations, and ultimately discourage successful collective action outcomes (Diamond, 2007; Rothstein, 2011). To what degree it is primarily input or output characteristics that generate policy failures and undesired outcomes remains an open empirical question. More and more countries are finding it hard to form governments capable of surviving in fragmented parliaments, and at the same time governing effectively (Rodrik, 2017). That market failure should be compounded by policy failure might appear as an evil but unlikely coincidence. Unfortunately, this is not so, because many market failures are related to the power and influence of some market agents. These agents will have power and influence not only in the market place but in the policy arena too. The dominance of perverse policies often reflects lobbying forces, characterized by large and powerful agents, and a political economy that systematically promotes them, particularly in the climate area (Berkes, 2006; Hepburn, 2010; Oreskes and Conway, 2010). The increased concentration of wealth and power of individual persons and companies may threaten the very fabric and functioning of democracy. The descriptions of market and policy failure above should not be taken to imply that the environment is constantly deteriorating everywhere and in all aspects. Many aspects of the environment are deteriorating while others have improved over time. We used to have smog in London that killed thousands and the air is today much

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better. There are famous incidences of dramatic pollution in the past that led to the creation of the EPA and the Clean Air Act and Clean Water Act in the USA and it is often asserted that many of these problems are solved over time. Interestingly though, Smith and Wolloh (2012) find that water quality across water bodies in the USA has actually not improved in the 40 years since the act. Even in those cases where local environmental quality has improved it may sometimes be at the expense of moving pollution elsewhere (symbolized by policies of building higher chimneys). See for instance Henry and Tubiana (2017) for one assessment.

3 THE MENU OF INSTRUMENTS Having identified some causes of ecosystem mismanagement, we turn to the menu of policy instruments available. Traditionally, policy making has occurred at the national level simply because this is where the power to implement and enforce exists. Increasingly, however, the powers and responsibilities for managing environmental resources such as forests and fisheries are being devolved to sub-national levels, in parallel with a recognition and acceptance of supranational policies required for managing global commons. One simplified way of organizing our menu of traditional policy instruments (adapted from Sterner and Coria, 2012) is in the overarching categories of 1) Pricetype; 2) Rights; 3) Quantity type regulation; and 4) Informational/Legal. The list below is by no means perfect or definitive. Many instruments can be classified in several ways. For example, tradeable permits could be considered either property rights or quantity type regulation. 1) Price-type • • • • •

Taxes Subsidy and subsidy-reduction Charges, fees or tariffs Deposit-refund Refunded charge

2) Rights • • • •

Property rights Tradable permits and quotas (Green) Certificates Common property resource management

3) Quantity type regulation • Technological standard • Performance standard • Ban

ARTICLE IN PRESS 3 The Menu of Instruments

• Permit • Zoning 4) Information/Legal • • • • •

Public participation Information disclosure Voluntary agreement Liability and financial rules International treaties

3.1 PRICE-TYPE INSTRUMENTS Our first category, price-type instruments, comprises those instruments that are designed to act directly on prices, such as taxes, fees, and subsidies. Whilst many instruments will ultimately affect market prices, what is unique about the instruments in the first group is that they directly add to or subtract from prices. That is, the instruments create incentives to change behavior directly through price, and thus “the market”, in such a manner and degree that will theoretically result in the cost effective allocation of the overall control burden (Hahn and Stavins, 1991). Sometimes the terms “market-based” or “economic” instruments are used to describe these policy instruments that we refer to as “price type”. We prefer to avoid this nomenclature, as the terms “economic” and “market based” can be both too broad (perhaps covering any rule recommended by economic theory) and too fuzzy (for example, covering cap and trade, which is primarily driven by assignment of property rights). The most famous of these price-type instruments is the environmental tax, often referred to as Pigouvian after the economist Pigou. A typical tax is levied by the state, with its proceeds going to the general budget. The purpose of a Pigouvian tax is to “internalize” the external effects. Such a tax has several effects: it makes the good or input which bears the tax more expensive, leading firms and households to economize on its use, whether through the use of substitutes, through changed technology, or changed consumption. The tax makes inputs to production, such as lead, sulfur or carbon emissions, more expensive. This reduces emissions directly, and also makes the final consumption good (comfort, miles traveled, lighting, etc.) more expensive, which affects consumption patterns (and thus again emissions – but this time indirectly). A Pigouvian tax is set equal to marginal damages, such that consumption is changed sufficiently to align private choices with socially optimal choices. The third effect of the tax is that it also gives the public sector revenues that can be used to reduce other taxes or increase public spending. The tax can be shown (under simple and ideal conditions) to be the optimal instrument, allowing other instruments to be benchmarked against it. The advantage that the same instrument serves several purposes (correcting for externalities as well as generating state revenues) has given rise to a whole literature on the so-called “double dividend”. A weak description of double dividend roughly states that an environmental tax has additional benefits. A strong (in fact,

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extreme) description states that environmental taxes have so many advantages that they are beneficial even without a primary environmental problem. It is an advantage if any environmental policy instrument also “happens” to provide tax revenue (of which many states are in chronic need). In general, taxes are hard to raise without distorting the economy away from optimal choices, whilst environmental taxes raise revenues and reduce distortions. The double dividend idea first appeared in the literature well before the 1990s, but it first received growing interest with early writings by Bovenberg (Bovenberg and Mooij, 1994a, 1994b; Bovenberg and de Mooij, 1997; Bovenberg and Goulder, 1996). It is also summarized well in Goulder (1995). Since these early writings on the double dividend, there has been much debate about its validity. The results depend very clearly on the exact question posed and exact baseline examined; see e.g. Jaeger (2011, 2012). A general conclusion across the literature examining the double dividend is that revenue recycling – or the use of environmental tax revenues to finance reductions in preexisting distortionary revenue – is real; see e.g. Goulder et al. (1997); Fullerton (1997). Returning revenues from market-based instruments as lower distortionary taxes rather than as lump sums raises aggregate welfare. The introduction of an environmental tax may cause the tax base to change due to the indirect effects on individual preferences. For example, West and Williams (2007) find that gasoline and leisure are complements. Raising the price of gasoline might therefore increase labor supply (and consumption), causing a fiscal welfare gain in addition to an environmental welfare gain. However, market-based instruments that raise little or no revenue are often favored for reasons of political acceptability to interest groups (Pezzey and Park, 1998). Double dividend enthusiasts sometimes argue that GDP would grow faster, unemployment would go down, and so forth, if only more environmental taxes were levied. In general, while the stronger versions are “too good to be true”, the more careful weak version – that Pigouvian taxes can have some other advantages such as revenue recycling – is quite reasonable. The counterpart of a tax is a subsidy, typically paid for by a nation state, which can be thought of as a negative tax. From an environmental perspective, conceptually, subsidies are used to encourage more of an activity that supports the environment and is under-provided relative to the social optimum. One common example is the provision of subsidies for abatement equipment in industry. Another is payments for ecosystem services (PES) when, for instance, farmers are paid to keep some attributes of nature in a certain state that might variously be beneficial for water fowl or endangered species (Wunder, 2005; Engel et al., 2008). In practice, subsidies tend to encourage the formation of lobbies by beneficiaries that serve to protect and prolong the subsidies (Baylis et al., 2008). In some cases subsidies have existed for so long that their original objective is gone or forgotten and they continue just because the lobbies defend them. This is so common that “subsidy removal” – the removal of perverse subsidies for inputs that are polluting (such as coal) – has come to be considered a policy instrument in its own right (Myers, 1998; Sterner and Köhlin, 2015).

ARTICLE IN PRESS 3 The Menu of Instruments

Subsidies have also been used by governments to encourage the development and/or adoption of particular existing or new technologies. The UK government has long subsidized nuclear power, whilst it recently canceled subsidies to support the development of carbon capture and storage. The US subsidies for carbon capture and storage technology have been praised by some, but criticized by others for promoting the continued use of fossil fuels (Conniff, 2018). Governments may also directly invest in, or subsidize, investments into adaptation and building resilience. Taxes and subsidies can also be levied by a municipality or agency, in which case a tax is often referred to as a charge or a fee. It is also possible, and increasingly popular, to design policy instruments that are combinations of fee and subsidy. These can be designed to be (approximately) revenue neutral. An example of this is refunded emission fees, where the proceeds of the pollution charge are not paid into the treasury or general budget but are earmarked for environmental purposes or repaid in some way to the general public, to the polluters, or to the victims of pollution. Section 5.4 gives more details of refunded emission payments in various countries. Other examples include deposit refund systems, tax, and dividend (for example, in the USA) and the Bonus Malus system (for example, for vehicles in France, where buyers of cleaner-than-average cars get a subsidy financed by an extra charge on the more emissions-intensive cars). Such combinations of fees and subsidies can use one instrument to influence behavior, and the second to compensate on aggregate for the cost of the first (see e.g. Section 5.4).

3.2 RIGHTS-BASED POLICIES The second category addresses rights-based policies. Property rights are the most fundamental of all rights-oriented “instruments” or institutions. Much of what is referred to as environmental policies can really be seen as a clarification of property rights – be they private, state, or communal. In the theory of rights-based management, researchers go into considerable depth concerning the formation and characteristics of rights (see for example, Leach et al., 1999, who give examples from India, South Africa, and Ghana). Property rights are often said to be a bundle of different rights, including notably the right to enjoy the use of a resource, but also in many cases the right to sell, lease, inherit or even destroy the property (Schlager and Ostrom, 1992). Thus, questions to be answered might include, for example, whether a land owner also holds rights to subsoil minerals and water, and whether she owns the rights to wildlife, radio frequencies, air, water, ecosystems, and space immediately adjacent to her land. When no one owns a resource (or when no one enforces their rights over a resource), market failures are likely to occur (Hanemann, 2014). Instruments that can proxy for property rights include tradable permits for emissions (for which new markets are typically created), and quotas for fishing or other resource use (which require the creation of markets only when the quotas are transferable). Tradable permits have even been proposed for regulating road transport externalities (Verhoef et al., 1997). Rights may be shared or held in common. There is a broad literature on the management of common pool resources, building on the

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work of Elinor Ostrom and many others, which shows that there are categories of resources which are particularly apt to being managed jointly, under common property regimes, CPR; see Ostrom (1990, 2000, 2005); Cole (2002); and Agrawal (2007). CPR may be a particularly appropriate instrument when, for instance, the economic benefits are sufficiently low and sufficiently variable as to make the costs of enforcing private property (e.g. division and fencing) unwarranted. Payments for Ecosystem Services (PES), as mentioned above, are often voluntary agreements in which those who benefit from ecosystem services pay those who provide the specific services. A water utility company might pay landowners to manage their farmland in the watershed so as to provide clean water downstream (Wunder and Wertz-Kanounnikoff, 2009). The rights-based analogue of this instrument is to create quasi property rights such as offsets or credits for some biodiversity, wildlife, water quality or forest cover attribute. Oil companies or developers can be required to buy a quantity of such credits that corresponds to what their business destroys. This creates a market and thus an opportunity for local communities to sell such credits.

3.3 REGULATION The third category, regulation, forms the core of environmental policy in most countries. Regulations directly control quantities of production or pollution, instead of the indirect approach taken with the price-based instruments. Activities may be banned totally or regulated so that they must be performed in a certain way, at a certain time, or within a certain area (for example, zoning forms the basis for much of city planning). This group of instruments is a lot less monolithic than one might think, and one should be wary of the simplified view that “economic” instruments are more flexible than regulation – often referred to as “Command and Control” (which does somewhat conjure up a Stalinist bureaucracy). Compare, for instance, a technology standard (which prescribes exactly what the firm must do) with a performance standard (which only specifies the environmental result – leaving the choice of technology to producers). The performance standard might be a certain maximum vector of emission coefficients in relation to output. It gives more freedom to entrepreneurs, which is valuable because producers often have information on technology options that a regulator might lack. It also somewhat helps to give the producers incentives to look for improvements themselves, i.e. incentives for research and development above and beyond what may be offered by alternative policies such as permits (see, for example Montero, 2002). However, as long as there is heterogeneity in pollution abatement costs, the standard will fall short of maximizing social welfare (which would require either pollution pricing by taxation or permit trading), see for example, Jung et al. (1996), Goulder et al. (1999). A technology standard, on the other hand, might tell the producer to use a particular kind of equipment or process. Technology standards are easier to inspect and monitor – it is easier to verify whether a firm has installed a piece of equipment than to measure its emissions. This is presumably the reason why technology standards (such as the height of chimneys, the existence of certain

ARTICLE IN PRESS 3 The Menu of Instruments

filters, or catalytic converters) are still quite common in legislation in many countries, in the reality of a second-best world. Even so, numerous scandals such as the Volkswagen emissions scandal in 2015, show how sophisticated firms can be in trying to circumvent monitoring and control efforts.

3.4 INFORMATION OR LEGAL-BASED POLICIES The fourth category contains other instruments that act by providing information, adjusting legal liability, or setting rules in other sectors such as the financial sector. Obvious examples include information disclosure rules. We take the approach here that it can be useful to cast the net a little wider and see that there are many rules in, say, banking and insurance that have considerable bearing on the environment or, for instance, the transition to a low-carbon economy. To take but one example, regulation of banking, insurance or even criminal liability clearly has effects on other productive sectors. In some cases, these effects may systematically affect technology choices. In the energy sector, for example, renewable energy can be produced at a much smaller scale than nuclear or other conventional technologies. Households may be a viable producer group but this requires a total overhaul of the requirements for information, the rules concerning liability, the writing of contracts and the way assets are calculated in energy companies, the responsibility for back-up power, the way electricity tariffs are structured or the manner in which creditworthiness is calculated by banks or other credit institutions, and so forth. Banking rules can thereby become instruments for sustainable management. Basel III, for example, is a package of reform measures, developed by the Basel Committee on Banking Supervision, to strengthen the resilience of the banking sector with respect to risk management. The package aims to improve the sector’s ability to adjust to shocks arising from financial and economic stress and to improve governance and risk management.1 Basel III was developed largely in reaction to the magnitude of risks in this sector. However, closer inspection shows it can also have important environmental effects. Because solar energy is more capital intensive (i.e., has a higher share of fixed rather than variable costs) as compared to fossil energy plants, different requirements on bank liquidity that change capital costs relative to labor or raw material costs may reduce the incentives to substitute solar power for fossil-based power production (Lowder, 2012; Gursoy, 2016). The profitability of different types of energy also hinges decisively on rules concerning the social cost of carbon and the discount rates that are applied to projects with public funding. Thus the rules or guidelines for discounting can turn out to be important policy instruments. The development and performance of the insurance sector can be very important for the assessment of damages from climate change and

1 See https://www.bis.org/bcbs/basel3.htm for more information.

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thus indirectly for adaptation and mitigation strategies. Insurance rules can also affect the cost of different energy sources or technologies (Mills, 2005).

3.5 THE PROCESS OF POLICY MAKING AT NATIONAL OR OTHER LEVELS The traditional approach to policy making focuses almost exclusively on instruments at the disposal of national governments, ignoring other agents and levels. Yet nation states are not the only agents to have a policy agenda. Municipalities, subnational regions, firms, NGOs, multi-national entities such as the EU or free-trade areas, and regional groups such as ASEAN or even the United Nations have their own priorities and agendas. The policy instruments they need or can use will often be somewhat different from those at the national decision makers’ disposal. NGOs or consumers’ groups might use labeling, education or information disclosure (Caswell and Mojduszka, 1996). Multinational alliances and the global community will typically have recourse to international treaties. This implies multi-tiered policy making, from the international to national to regional arenas. An added complexity is that some instruments at, say, the local level might be more suitable to combinations with certain instruments at the international level. As an example, a global cap and trade system is hard to combine with other local instruments such as taxes or technology subsidies. On the other hand sometimes companion policies exist and need to be dealt with carefully; see for instance Coria (2009) and Burtraw et al. (2018). The rise of corporate environmentalism (see, for example, Hoffman, 2001) – and its counterpart, the growth and globalization of environmental NGOs – has resulted in a diverse group of non-government actors searching for policy instruments. Large multinational corporations have agendas that may be very relevant for the environment, yet their background may be quite complex. For example, fossil firms may be expected to be less favorable to addressing climate change than firms in renewable energy. However, over time the former may see themselves more as “energy” companies, and with technical progress they can transform their business models (Pulver, 2007). Private sector organizations may be more science-based and more farsighted than many political organizations driven by the electoral cycle. Firms that are high users of fossil fuels, or high polluters, most likely understand that they will be regulated and are mainly interested in affecting the timing, the methods and the international coordination of regulation that is inevitable, and thus they will be strategic. Environmental NGOs, such as WWF, Greenpeace, EDF or the Nature Conservancy, have millions of members and large professional staffs and are quite influential, particularly in the US. The process of policy making, with its risks of policy capture, distributional and equity effects of instruments or the opportunity created for political corruption has been discussed in the literature for some time (see Olson, 1965; Stigler, 1971; Posner, 1974; Peltzman, 1976; and Becker, 1983). Policy capture reflects the tendency of powerful groups to steer policy away from the social optimum. A distributional question in environmental policy is whether the poorest nations or people will

ARTICLE IN PRESS 4 The Selection of Instruments

be losers or gainers. Corruption includes misdirection of resources that were intended for the public good. Just as behavioral economics looks at individual behavior, political economy studies how political institutions adopt, implement, and enforce policies, including environmental policies. Keohane et al. (1998), for example, presents an explanation of why the environmental policy instruments selected in the United States for decades have diverged strikingly from the recommendations of normative economic theory. Often when policies are badly needed to counteract the dominance of strong economic players, we see there is a risk that these same players will be able to corrupt the policy process in their favor. Sometimes this implies that policy failure is added on top of market failure. Many of the growing environmental problems are complex, trans-boundary global issues, which are likely to require new approaches that do not fit in the traditional policy instrument columns above. New environmental policies are likely to be needed to initiate cultural transformations, new lifestyles or consumption habits, technological revolutions and overarching sustainability policies. Such policies will most likely need to reach across traditional boundaries, regarding the creation of new institutions or population policy to achieve an optimal or sustainable scale of the economy. Some observers call for stopping growth or engaging in “de-growth”. Though literally degrowth is the opposite of growth, a decrease in production, the term is often intended not as an economic concept, but signifies more of a social movement that calls for the downscaling of production and consumption in higher-income countries (Demaria et al., 2013). From an economist’s perspective, if growth is sensibly defined to mean an increase in that which is desirable, which includes taking into account ecological restrictions, then growth should be consistent with aspirations to improve human wellbeing, such as through improving health or education services, or access to clean water, without threatening ecological systems. The relevant problem then becomes defining what sustainable growth is and to find the policy instruments that promote that kind of growth while minimizing growth of sectors or technologies that have negative repercussions.

4 THE SELECTION OF INSTRUMENTS Having looked at market failures and at the menu of policies, we turn now to policy selection. To derive theoretical results in economics, we are generally obliged to simplify and abstract from numerous real-world circumstances in order to focus on one particular aspect of the market failure. The optimal choice of policy instrument is likely to be affected by market structures, existing market failures, transactions costs (see the review article by Krutilla and Krause, 2011), asymmetric information, considerations of equity, and preferred trajectories over time. Some of these aspects have been the object of considerable analytic effort and we are thus able to repeat some clear-cut results that are central to the profession (see Stavins, 1995, 2003; and Sterner and Coria, 2012, for a fuller treatment).

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4.1 EFFICIENCY At the heart of economic theory on policy selection is a consideration of economic efficiency. We draw here upon some of the most basic lessons from the functioning of the overall economy. Competition between different producers in a market tends to produce an outcome that is “economic” in the direct sense of the word: it saves resources. The intuition behind this is that competition will spur effort and innovation (both technological and organizational). This will reward the best producers, which will tend to be favored by competition. Adam Smith praised the competitive market precisely because it is the profit motive and not altruism that makes sure we get the best bread from the collective of bakers (Smith, 1776). In a similar vein, environmental economists like to point out that we could get a cleaner environment at a much lower cost if we used general instruments that build on market principles rather than clumsy regulation. The subject of economic efficiency in environmental instrument choice deserves careful consideration. Regulations are often less flexible than would be optimal. For example, if a total reduction in pollution is required, regulators might require equal reductions from each source of pollution, rather than taking into account that abatement costs often are heterogeneous. That is, different producers will have different marginal costs of achieving the same reduction in pollution in which case, the usual market analogy applies. Cleanup costs could be reduced by allowing the best performers in the market to do more of the “heavy lifting”. The greater the heterogeneity in costs, the greater the saving from exploiting this difference. For example, if abatement in one firm is half as costly as in the other, a society can save around 10% of total costs of cleanup by allowing the firms to distribute the abatement cost efficiently than if they were constrained to do the same amount each. If one firm can undertake abatement at one-tenth of the cost of the other (i.e. 90% cheaper), then savings of around two-thirds of the overall costs can be achieved through market allocation (for details see Sterner and Coria, 2012, Table 9.1, p. 146). Next, we need to consider what types of instrument will lead to this efficient market allocation. A key requirement is the existence of a price on the pollutant. This could be achieved by using a subsidy, a tax or a physical regulation such as a certain abatement requirement, so long as this requirement is tradable. A tax will incentivize those firms with lower abatement costs to do a larger amount of abatement. A regulation provides a similar incentive – if trading of permits is allowed – because then the company where abatement is cheaper will offer to do the abatement, selling its “right” to pollute to the plants where such abatement is more expensive, and resulting in the same total abatement but at lower cost. Cost heterogeneity is not the only factor. The degree of competition is also important. In the extreme case of a single monopolist, the cost of abatement will be the same for any given level of abatement irrespective of instrument used. However, the performance of taxes (or other price type instruments in first group, above) in the face of monopoly or restricted competition is often quite complex. One reason for this is that monopolists already tend to under-produce (not pollution but products). A tax will tend to aggravate this problem. If we have a polluting monopolist, we are in fact

ARTICLE IN PRESS 4 The Selection of Instruments

faced with two market failures: the monopoly and the externality. This will typically require two instruments: one to increase competition and one to reduce pollution. The case of a monopoly is relatively simple, but that of duopolies or other market structures can be analytically very complex to deal with (Simpson, 1995). Studying an interconnected web of countries with trade complicates things further. Whether standards or taxes dominate has been shown theoretically to depend on, inter alia, whether trade is modeled as a Stackelberg or Cournot equilibrium (Ulph, 1992). More generally, when there is imperfect competition in global markets, pollution taxes can lead to “pollution shifting” and “rent capture” (Kennedy, 1994), and governments may relax or strengthen environmental standards, depending on specific market structures (Barrett, 1994). Environmental regulations in large countries will affect world prices of traded commodities, suggesting the need for second best environmental taxes (Krutilla, 1991). Subsidies and taxes may look similar in the simplest of models. Both raise the relative cost of pollution and lower that of abatement. If production is clean, the firm either avoids paying a tax or it receives a subsidy. At the margin, the last unit of pollution will cost the firm $x either way – either as a tax or a subsidy lost. In a more complete analysis, however, the tax and subsidy are clearly not identical. A company that is on the very verge of bankruptcy might be pushed over the edge by an environmental tax scheme but it might be saved by a cleanup subsidy scheme. We can (again under certain simplifying assumptions) show that the correct and optimal instrument is the tax, and not the subsidy. The final product price should include the shadow rent on scarce environmental resources and companies that do not bear their total costs should go bankrupt. But of course these arguments are hard to make in the face of angry lobbyists or politicians who are worried about job loss. We will come back to political feasibility and similar arguments later.

4.2 INFORMATION ASYMMETRIES AND UNCERTAINTY Another area that restricts instrument choice is the character and availability of information flows and the structure of incentives. Typically high levels of uncertainly exist about the environmental damages themselves. For example, even when a pollutant is known to cause adverse health or environmental damage, the exact extent of that damage may be unknown. Though climate change is known to be a significant global challenge, substantial uncertainty surrounds its potential economic damages. Toman and Shogren (2010) provide a broad assessment of uncertainty surrounding costs and benefits with respect to climate change and the relevance for policy. See Maler and Fisher (2005) and Aldy and Viscusi (2014) for a discussion of risk and uncertainty in the environmental realm. Here, we focus on uncertainties specific to policy-making: 1) asymmetries between local agents and central authorities; 2) uncertainties/information asymmetries regarding abatement costs; and 3) uncertainties that arise when pollution cannot be readily measured or monitored. Typically, in an economy, the agents most directly involved have better information than the central authorities when it comes to local costs of technology and

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resources. The dilemma is that those best informed may have an incentive that is not aligned with social preferences. It is important that an instrument mobilizes these agents, and indeed the advantage of decentralization lies in empowering local agents to use their knowledge and giving them incentives to search for better means of production.2 There is a huge literature that analyzes the often complex instruments that policymakers need to design to entice polluters to reveal information to them through their choices of pollution management contracts (or water, service provision, etc. contracts). See, for instance, Laffont and Tirole (1993), Hoel (1998). Uncertainty about abatement costs is the epitome of a well-studied special condition in environmental economics. Weitzman (1974) shows that, for a special type of uncertainty concerning abatement costs, both quantity-type regulations and taxes may lead to certain errors, because the abatement achieved will typically not be optimal from the viewpoint of the ex post knowledge about costs. However, the expected cost implied by the deviation will be different for standards versus taxes. If we know the slopes of the costs and damage functions, then it can be shown that price-type policies are preferred on social welfare grounds if the abatement costs are steeper, and quantity-type regulation is preferred if the damage costs are steeper. This is a result that has sparked a good deal of further research that looks at more complicated or realistic cases (e.g. Pizer, 2002), for example, if there is a correlation between benefits and costs (Stavins, 1996). A particular kind of uncertainty occurs when pollution cannot be readily measured or monitored, for instance, Non-Point Source Pollution (NPSP). Examples are found in run-off of nutrients such as phosphorous and nitrogen in agriculture (see Carpenter et al., 1998, for one such example in the literature). Government authorities may observe aggregate pollution in, for instance, a body of water, but may not have the capability of measuring and assigning individual responsibilities to individual farmers. This again has spurred a large literature looking at different situations. For example Xepapadeas (1992) addresses policy instruments for dealing with NPSP when there is pollutant accumulation, thus requiring a dynamic framework. Cabe and Herriges (1992) use a Bayesian approach, addressing asymmetric information and transport costs. One line of thinking builds on the assumption that, though the central authorities are incapable of monitoring, it may be possible for the decentralized agents themselves (the farmers) to monitor each other. Sometimes it may be possible to create a system where the local agents have both the incentive and the capability to jointly manage a common pool resource (Ostrom, 1990; Cason and Gangadharan, 2013).

2 These arguments do generally speak in favor of using a price mechanism but it is important to recognize that, for most environmental management, the bread and butter of environmental protection agencies is often regulation. It is worth considering why this is so, but the fact remains that regulations are frequently used and some regulations are much better than others from the viewpoint of allowing agents a greater freedom of choice, as was discussed in Section 2 concerning the difference between technology and performance standards.

ARTICLE IN PRESS 4 The Selection of Instruments

4.3 INTERTEMPORAL EFFICIENCY Dynamic efficiency – that is, efficiency of resource use over time – is a particular area of study often related to resource scarcity (Nordhaus, 1974). A given resource may be fixed or it may only grow at a certain rate under certain conditions. The body of work known as natural resource management deals with how to allocate use over time. At first glance, it may seem impossible to allocate a finite resource over infinite time. Conceptually, however, it is quite easy. If we use only a fraction of the remaining resource each year, it can “last” for centuries on the condition that we use less (in absolute terms) each year. This might occur in parallel with gradually improving technology or successive improvements in substitute goods or inputs. If so, one can also show that such a resource allocation will typically be associated with exponentially rising scarcity rent, so that the price of the resource rises, encouraging successively more frugal consumption each year or the use of more effective production technologies. Hotelling (1931) showed that such a regime with consecutively rising scarcity rents is to be expected (under certain ideal assumptions) if there are secure property rights to the resource in question. (Variations may be due to all kinds of different assumptions with respect to market structure, technology and so forth.) There are various ways to prove or illustrate the Hotelling principle. The relevant conditions emanate from mathematical optimization but the most intuitive economic explanation comes from the so-called arbitrage condition.3 The modern field of intertemporal resource and environmental economics was pioneered by a series of articles by Geoffrey Heal and Partha Dasgupta that is nicely summarized in their seminal book, Dasgupta and Heal (1979). In policy-making, Hotelling insights apply to the intertemporal performance of regulations as well as cap-and-trade programs with banking (see e.g. Rubin, 1996; Wigley et al., 1996; Kling and Rubin, 1997). A crucial issue for policy making is how to deal with those resources that lack secure, or indeed any, ownership rights – such as the atmosphere. In an analogy with Hotelling, we might expect that there should be a policy instrument such as a scarcity rent on carbon and other greenhouse gases in the atmosphere and that this scarcity rent should rise analogously with a Hotelling rent. Interestingly, there is a large literature analyzing what happens in other cases. The Green Paradox literature (Jensen et al., 2015; Sinn, 2015) shows that perverse incentives can be created. If, for instance, the policymaker announces that in the year 2030 there will be a high and rapidly rising carbon tax, she may inadvertently be creating an incentive to produce and burn as much coal as possible today (before this policy depresses the rents to be earned).

3 Simply stated, an investor who invests X $ in a business, some stock, foreign exchange etc., would

normally expect these assets to have a total value (including dividends, and correcting for uncertainty) in the future equal to X(1 + r)t . If one asset is not expected to rise in value so that its future value is equal to its current value times (1 + r)t then it must fall in value immediately otherwise no one would buy it. Thus the spot price (or actually scarcity rent) of that asset falls immediately and when it has found its correct value then it will again be expected to rise exponentially.

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When resources are biological and their growth dependent on temporal, spatial and ecological constraints, we refer to the area as natural resource economics (see textbooks by Fisher, 1981; Conrad and Clark, 1987; Hanley et al., 1997; or Conrad, 1999; and Brown, 2000, on the characteristics and management of renewable natural resources). When multiple ecosystem services related to the natural resources are taken into account, this area of research becomes particularly rich and interesting, with variations or interdependencies between growth rates of various species, possibly featuring nonlinearities, thresholds and complex tradeoffs between different sustainability goals4 (Dasgupta, 1983; Brock and Xepapadeas, 2010).

4.4 SPATIAL EFFICIENCY Just as the benefits of environmental goods and the costs of negative externalities are not spread evenly across time, they are typically not spread evenly across space. This spatial heterogeneity may be driven not just by a particular environmental good or a particular externality, but also by the location of people who use an environmental good or amenity. In either case, policies that are applied uniformly across space may be suboptimal, and spatial targeting may provide better outcomes (Sanchirico and Wilen, 2005; Albers et al., 2017). Just as we need to optimize resource use over time, we also need to optimize the spatial allocation of resource use. As an example, at the most basic level, if a remote area of forest is far enough from human populations that it is protected by distance alone, it would be inefficient to allocate scarce enforcement budget to that area. Conversely, areas of forest very close to human populations may be very expensive to protect and thus it might again be inefficient to enforce there, but for very different reasons (Albers, 2010). Thus, one might not be surprised to see zoning of protected areas and forests, particularly in lower-income countries where people are highly dependent on the resource base, in which more distant areas do not need to be protected; very close areas are too expensive to protect; and so protection is concentrated at intermediate distances from villages or towns, with a possible buffer zone adjacent to the community (Robinson et al., 2013). As another example, payments for ecosystem services can often times be more effective when spatial aspects (the locations of costs, benefits, and those affected) are taken into account. This includes environmental service provision, risk of environmental service loss (e.g. through deforestation), and the cost of participation for the landowners (Wünscher et al., 2008). Similar arguments apply to the management of fisheries and marine protected areas, with the additional spatial element that most fish species move across institutional boundaries. Similarly, “Marine Spatial Planning” requires “appropriate planning activity at different spatial scales” (Gilliland and Laffoley, 2008, p. 787). In the US and Canada, policies that have been proposed to encourage soil carbon sequestration have been demonstrated to 4 As an example, with respect to forests, optimal rotations tend to be different than the Faustmann rotation age when carbon sequestration and other ecological, non-market ecosystem services are taken into account.

ARTICLE IN PRESS 4 The Selection of Instruments

be inefficient because they do not take into account spatial heterogeneity with respect to bio-physical and economic conditions (Antle et al., 2003). Yet measuring site specific differences introduces additional costs (Wu and Boggess, 1999; Antle et al., 2003). GIS-based spatial analysis is enabling improved spatial targeting for policy interventions such as agri-environmental schemes (van der Horst, 2007), and may make spatially targeted policies more cost effective. Biodiversity offsets are an explicitly spatial rights-based instrument (see Section 3 above), which allow developers to build at locations where biodiversity or habitat more generally will be negatively affected, so long as the ecological damage can be offset elsewhere, resulting in “no net loss” or a “net gain” (McKenney and Kiesecker, 2010). Biodiversity offset schemes have the potential to encourage developers to build in a different location, or in a different way, rather than undertake the offset. Thus, offsets can complement other regulatory arrangements such as standards or pricing approaches (taxes and subsidies), and can be considered closely related to tradable permit instruments. However, offsets can only be employed where the habitat or biodiversity being considered is spatially substitutable. As such, “critical capital” ecosystems that cannot be substituted could not be included. Examples of biodiversity offsets include the Wetland Mitigation Banking scheme, in the US, and many in South Africa that fall under the Western Cape Provincial Guidelines on Biodiversity offsets. Legislation is required to set out roles and responsibilities of developers, regulators, and groups representing third parties (Groom et al., 2014). Biodiversity offsets can be controversial due to the idea that biodiversity and habitats and locations in one area can be substituted elsewhere, and thus such ecosystem services are fungible. Habitats may be reduced to overly simplistic metrics to enable spatial substitutability, and the local benefits that are provided by these habitats for nearby communities may not be fully recognized (Bateman et al., 2013).

4.5 PRACTICAL AND POLITICAL ASPECTS In many practical applications, instruments such as individual pollution limits or pollution taxes are hard to use due to lack of adequate information, and other instruments must be designed. Often we have a multiplicity of these conditions at the same time and the complexity may be such as to defy rigorous analysis, turning instrument selection and design into something of an art. Considerations of lobbying and power inevitably lead one to think of political economy and of political process. The terminology of instrument choice may then be perceived as a social engineering approach. A complementary approach is to think of designing a process in which parties (be they polluters and victims of pollution or be they countries) negotiate a policy and the appropriate instruments of implementation and enforcement. Climate policies, for example, can either be discussed in terms of designing tax or cap-and-trade systems on the one hand, or, on the other hand, the design of international climate negotiations. In either case distributional problems are paramount, either within a country or among countries. Partly, the appropriate instruments and processes depend on the characteristics of the problem at hand.

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The process of political decision making in the real world tends to include difficult and sometimes not very transparent trade-offs between various interest groups. Environmental interests often argue for the “polluter pays” principle, which holds that polluters should not be able to pass through the cost of abatement to others. Conversely, existing polluters lobby to be “grandfathered in” and allowed to continue as before (Markussen and Svendsen, 2005; Lockie, 2013). Woerdman et al. (2008) address efficiency and equity aspects of each approach. Another complication is that the design of policy instruments should also take into account ecological realities very far away and on very different systems than those they originate from, due to complex natural transport and transformation processes. This is likely to make it more difficult to apply good regulation when polluters and polluted never physically meet. In such circumstances, negotiations between the different parties could be more difficult as well. In addition, some ecosystem processes tend to unfold slowly and delayed effects of actions taken in the past may suddenly change the nature of the problem. Using an approach that from the start integrates biosphere and social characteristics should improve the chance of success of a policy response simply by being able to account for more contingencies than when only taking a social perspective (Levin et al., 2013). In light of the catastrophic harm that is likely if planetary boundaries are transgressed, the “precautionary principle”5 argues that great efforts are necessary to minimize that risk. When processes are complex informationally, spatially, and socially because they involve many stakeholders, there is a risk that analysts are invited in only after the main policy choices have been selected. Then economic analysis is only used to compare a small number of given scenarios, for example, options for a new road, instead of being asked to compare a whole menu of solutions to an urban traffic problem including congestion pricing rather than roadbuilding. This is of course quite unfortunate. The discussion above on policy instruments is focused on instruments needed to reach an ideal state, to maximize welfare under market failures. It presupposes strong powers for the decision makers. In reality there are limits to political influence. Resistance to government is rife – and increasing in many parts of the world. There is furthermore often a lack of institutions – an absence of governance at the appropriate (local) or global levels. Finally, it is naïve to assume that governments simply seek to maximize welfare. They have their own interests, especially the incentive to be re-elected (or survive revolutionary attempts). 5 The precautionary principle, which started off as an environmental concept to deal with “increasingly

unpredictable, uncertain, and unquantifiable but possibly catastrophic risks”, through pre-damage rather than post-damage control, has morphed into a concept that incorporates an ethical stance (Comest, 2005: p. 7). It has been defined as follows: “When human activities may lead to morally unacceptable harm that is scientifically plausible but uncertain, actions shall be taken to avoid or diminish that harm” (www.precautionaryprinciple.eu/). See Gollier et al. (2000, p. 245) for an economic interpretation of the Precautionary Principle, which concludes “that more scientific uncertainty as to the distribution of a future risk – that is, a larger variability of beliefs – should induce Society to take stronger prevention measures today.”

ARTICLE IN PRESS 4 The Selection of Instruments

Instead of assuming that governments strive to maximize social welfare and need advice on how best to do so, we might take an alternative worldview, seeing governments either as representatives for various economic interests or as autonomous agents maximizing their own utility. This may involve promoting the interests of powerful lobbies or it may be conceived of as selling “services” to such interests in return for support or simply money to be used to win elections. Either way, such models open up a much more active role for lobbyists and much larger risks of the capture of political processes by special interests.

4.6 NORMATIVE PRINCIPLES, DISTRIBUTIONAL ASPECTS, AND ENVIRONMENTAL JUSTICE Both environmental damage and associated abatement costs may differentially affect the rich and the poor or other significant groups in society, with equity implications across space and time. Whether they do so is often a very sensitive and hotly debated issue. Attention is often focused on the important cases in which the poor suffer the most from pollution. This is sometimes even used as an argument for more environmental policies. However, we can find empirical examples with different incidences of the distribution of the burden. Many policy instruments are hard to use because they have distributional effects that may have adverse effects on various segments of the population that command political influence (Barbier, 2011; Ward and Cao, 2012). Indeed, the same resource problem might be solved with different instruments, such as taxes, cap and trade, subsidies or regulations, with each typically having a different distributional effect. For example, different companies, social groups or countries will bear different proportions of the costs and benefits. Moreover, many of the agents of an economy are acutely aware of such distributional effects and thus real policy making is not a neutral technocratic choice of “optimal” instruments, but often better portrayed as a struggle between different economic forces, each of which is lobbying for some instrument or process they believe will be advantageous. Thus, issues of fairness and political process are crucial (Carlsson et al., 2013). This is particularly so when it comes to sharing risks and uncertainties, where the poor typically are more risk averse. That the poor will bear an unfair share of the costs of abatement has also been used as an argument against clean-up or abatement efforts. The Clean Air Act in the US, for example, was found to be regressive, because lower-income households allocate a larger share of total spending to energy, such as electricity and gasoline, that depends on fossil fuels (Gianessi et al., 1979; Fullerton, 2008). However, often there is equally a problem when a proposed instrument threatens the rich and powerful, who are effective at lobbying the political process. Some instruments are specifically designed to appease special interests by allocating favorable exceptions or emission permits to them. The truth is of course much more nuanced and hard to pin down, and it is useful to step back and consider who holds the relevant rights. It is further vital to consider how

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the presence of inequality at different scales (regions, countries, rural/urban, within urban areas, etc.) affects the way that policy instruments should be structured. Factors to be considered might include non-uniform/discriminatory pricing (taxes/subsidies); and different mechanisms of allocation of property rights, regulations, etc. Rights to nature may – much like land or water rights – be allocated to society, to the victims of pollution, or to current polluters (“grandfathering”). Property rights are essential for the definition of all other policy instruments (see Section 3). They appear to have developed in quite different ways in different countries. There are countries (such as the USA) where private property rights are very strong. With land rights may come the rights to minerals below the ground, hunting and fishing rights, rights to adjacent water bodies, and rights to build on the land with few or no restrictions. For a historical perspective on below the ground rights in the USA, see Libecap (1978). For a broader discussion of “3D” property rights, see Kitsakis and Dimopoulou (2014). There are other countries, including many in Europe and many lower-income countries, where the state keeps unto itself many of these rights (Heltberg, 2002). Historically, it seems that property rights developed first for immediate belongings and then for land. Water rights have a separate history but are sometimes related to land rights. The rights to minerals, to emit gases into the atmosphere, or to broadcast radio waves at different frequencies are examples of rights that have evolved over time as society has perceived a scarcity or a congestion that required such definitions. This process can be likened to the historic process by which land was privatized. In feudal societies like England or France, there was a process by which common lands were privatized that is often referred to as enclosure, because in England private land was often “enclosed” by fences or hedges (Wordie, 1983, provides a detailed discussion of enclosures in England). Different concepts of pollution rights can be seen as a counterpart corresponding to the distinction between different instruments such as taxes, subsidies and revenue neutral tax-subsidy or refunded tax schemes. If a polluting business is seen to have the rights to nature, it can be allocated property rights to land, to water or to pollute the air. The price-type instrument counterpart would be a subsidy: if a business owns the rights, then society can still require a cleaner environment – it just has to pay for it. On the other hand, if the rights reside with the state or the population of a country, then business must buy such rights. The corresponding price-type instrument is a tax,6 which means that firms can pollute – but they must pay for every unit of pollution. The quantity-type analogue is a tradeable permit scheme with auctioned permits, and, again, the cost is incurred by the firm. Table 1 provides a different view of how policy instruments can be categorized by their assumed property rights. In Column 1, polluters have absolute rights to the environment, while in Column 4 society (interpreted perhaps as a representative of the victims of pollution) has these rights. Columns 2–3 are intermediate. The table’s 6 Coase (1960) makes an important point when he says that the rights may be allocated either way – optimal

allocation can still be feasible. However, if the rights are unclear, and transactions costs sufficiently high, then it will be hard to reach an optimal allocation.

ARTICLE IN PRESS 5 Selected Examples

Table 1 Policy instruments and property rights Instrument

Ownership rights to the environment Polluter Mixed Victim or society (partial) (polluter pays principle) (1) (2) (3) (4) Public cleanup CAC, VA, TEP TEP, partly TEP, auctioned (free) auctioned Subsidies REP, Partly REP BM, Tax, DRS taxsubsidy TD Polluter (absolute)

Quantity-type Price-type

Notes: CAC = control-and-command policy; VA = voluntary agreement; TEPs = tradable emissions permits; REPs = refundable emissions permits; DRS = deposit–refund scheme; BM = bonus malus; TD = tax and dividend.

first row is for price-type instruments while the second is quantity-type. This helps us disentangle a confusion that is quite commonly made between “cap and trade”, which is a quantity-type instrument and often thought of as more “business-friendly”; and tax, which is a price-type instrument that generates revenues for the state and implicitly gives the state rights over the environment. However, it should be clear that the choice of a price vs quantity instrument can be made irrespective of the desired distributional effect. A price instrument need not be a tax. If society considers the polluters to have the rights to the environment, or if for political or pragmatic reasons a distributional effect that benefits business is desired, then subsidies can be used. If rights are intermediate, revenue neutral instruments such as Bonus Malus, refunded emission payments, or tax and dividend can be used (see Section 5.4). Similarly, with quantity type instruments, the rights can be auctioned or allocated for free depending on the balance of rights or of power and income distribution effect desired. Some instruments like the voluntary agreement seek to achieve a reasonable allocation without open coercion or payments although there is often an implicit threat of other instruments, see Carraro and Lévêque (2013).

5 SELECTED EXAMPLES Any particular collection of examples will always be in some sense arbitrary. We choose to start with a tax because it is often thought of as the prime example of an environmental policy instrument. Because climate change is at the same time such a dominant environmental problem, we start with carbon taxes (Pearce, 1991). We turn then to fuel taxes, which can be seen as a sectoral carbon tax. In Section 5.3, we look at cap and trade schemes as an alternative mechanism for pricing carbon and in Section 5.4 we look at the refunding of environmental fees. Section 5.5 looks at regulation versus pricing, taking industrial solvents as an example. In Section 5.6, we turn to the influence of culture on behavior, and in particular tobacco use and meat consumption.

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5.1 TAXING CARBON Despite being the poster child of environmental economics theory and despite some attractive properties, carbon taxes are not nearly as common a policy instrument as might be expected. To date, significant carbon taxes have been instituted mainly in a few Northern European countries and in the Canadian province of British Columbia.7 Recently, France – which has a long history of trying to implement carbon taxes unsuccessfully – has emerged as a very interesting example with a strong focus on carbon taxation. In 2018, the French tax reached €45/ton CO2 and by 2022 it is scheduled to reach €86, which will bring it quite close to the tax-level in Sweden. The reasons why it has been so hard to get significant carbon taxation in other countries are not clear, though it is possible that lobbying plays an important role, particularly in countries that have fossil industries of their own. What is clear, however, is that a carbon tax, unlike a narrower sectoral regulation, attracts more hostile lobbying from fossil fuel interests, which have a lot to lose (Bjertnæs and Fæhn, 2008; Blackman et al., 2010; Sterner and Coria, 2012). It is likely that access to good examples of successfully implemented carbon taxes is important – perhaps the recent example in France, a major European country, will help break some of the resistance. To try to cast more light on how a carbon tax can function, this section describes some of the experiences of the CO2 tax introduced in Sweden in 1991. The Swedish experience is of interest because it is by far the highest level of taxation (more than twice as high as the neighboring countries Norway and Finland that have also had high carbon taxes for a number of years) and has been applied broadly and seemingly painlessly in a modern economy; see Hammar et al. (2013). In 2017, Sweden’s general CO2 tax corresponded to €117 per ton CO2 ,8 which can be compared with current tradable permit prices within the EU emissions trading scheme (ETS) around €10 per ton. The development of CO2 taxation and the use of revenues are determined in accordance with general Swedish national budgetary rules. A central element is not to earmark tax revenues for particular purposes; instead, the spending of tax revenues is decided in the normal budget process. Throughout the existence of the CO2 tax, policymakers have aimed at ensuring a balanced tax design. The tax was introduced in a step-by-step manner starting in 1991 and undesired distributional consequences on low-income households have been addressed by adjustment of the come tax rules. Sweden has applied quite high taxation to energy carriers for a long time. Up until the 1970s, the primary reason for taxation was to raise public revenues; taxation consisted of a single energy tax. In 1991, there was a major tax reform and Sweden complemented the energy tax with specific CO2 and sulfur taxes, because environmental policy was becoming increasingly important in the political agenda. The CO2 7 We highlight significant cases of high carbon taxes here. In addition there are many interesting examples of carbon taxes (generally at lower levels) in other countries around the globe as well. For a recent update of all existing carbon taxes worldwide, see the World Bank’s annual ‘State and Trends of Carbon Pricing’ reports. 8 1130 SEK per ton.

ARTICLE IN PRESS 5 Selected Examples

tax was introduced on all major fossil fuels at rates equivalent to €27 per ton CO2 , in addition to a separate energy tax. As a result, although the energy tax was reduced, the combined energy and CO2 tax rose, which meant that fuel prices increased, and all fuels were taxed at very high levels compared to other countries. The introduction of this carbon tax was part of a major tax reform that included dramatically lower marginal income taxes on capital and labor, the elimination of various tax shelters, base broadening of the value-added tax, and major reform to property, wealth, inheritance, and corporate taxes. The political opportunity to introduce this rather unique tax consisted of the confluence of two separate political processes. On the one hand, there was demand for a drastic reduction in marginal income tax rates, which had reached very high levels (in some cases around 90%). The highest marginal tax rate was reduced to 50% (later it rose somewhat again but only moderately compared to the historic levels). At the same time, there was an increasing interest in environmental issues, politically and throughout Swedish society. The CO2 tax was thus introduced at a moment when there was a need to fill a gap created by reduced taxes on other factors of production. The tax yield from changes in energy-related taxation amounted to roughly one percent of GDP in 1991, of which introduction of value-added tax on energy consumption accounted for the major part. Importantly, the introduction of this new tax was not associated with an overall increase in the burden of taxation. On the contrary, the overall tax level in society, as measured by the tax share of GDP, although still high, fell substantially (by about ten percentage points) during the period when the carbon tax was introduced and raised. In other words, the effect of reducing other taxes was more important and the carbon tax thus helped finance other tax cuts – though only partially. The CO2 tax rates have been significantly increased over the years, with the purpose of achieving cost effective emissions reductions. The tax changes have, however, been implemented stepwise so that households and companies have had time to adapt. Typically, tax increases for companies and households in the energy and environmental areas in Sweden have been combined with general tax relief in other areas in order to avoid increases in the overall level of taxation, address undesirable distributional consequences, and stimulate job growth. Such a combination of measures has been the result of a desire to design the tax scheme in a way that ensures a sufficient balance between different policy considerations. Over the years, there has been a general consensus among the different political parties in Sweden to focus on the CO2 tax as the primary instrument to achieve greenhouse gas emission reductions. Sweden has had governments that were relatively left-wing and right-wing, but this has not led to any major deviations from the chosen road forward in this regard. Also, all major Government proposals are based on in-depth analysis and review by independent committees, experts and stakeholders. The advantage of a CO2 tax is that it is a market-based instrument, which enables households and firms to choose measures to reduce fossil fuel consumption – and thus greenhouse gas emissions – that are best suited for their specific situation. However, the effect of the CO2 tax was also complemented by aid schemes for limited

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time periods, to ensure that real options are available for households and firms. In Sweden, such schemes have included support for investments in energy savings in buildings, fossil-free electricity production, infrastructure projects for public transport, and urban district heating systems. Sweden is a fairly typical modern market economy so we think its experience of carbon taxation should generalize to other countries. It is worth reflecting however on whether there are any special features that helped make it possible to introduce such taxes. One striking factor is a long history of consensus politics and total absence of any fossil fuel industry – but on the other hand good access to hydropower and bioenergy resources. One factor does however need to be highlighted that applies to all “small open economies”: trade is a very high share of GDP (around 45%) and the economy is dominated by a small number of very large firms. There is therefore a well-founded fear of loss of jobs and of carbon leakage (when national policies targeting externalities locally lead to an increase in demand or movement of production elsewhere). For example, global net emissions could stay the same or increase if production moves from a country that is taxing carbon at a high rate to a country that is taxing carbon at a lower rate or not at all. To avoid these risks and make the policy more palatable to industry, Sweden only applies the full tax rate of the carbon tax to those sectors that are somewhat immobile or protected from full international competition. The idea is that when trading partners face the same tax levels, then Sweden will impose the full tax on sectors exposed to competition. Large industrial use of energy has therefore, ever since the introduction of the CO2 tax, faced a lower tax level (which has varied from 20 to 50% of the base level). Furthermore, most of industry is subject to the Europe-wide carbon emissions trading scheme. Initially, firms were subject to both carbon tax and the obligation to participate in the EU ETS. However, in 2011, the CO2 tax for industrial installations within the EU ETS was abolished, because companies complained of double taxation and because national policies for emissions subject to a cap at the EU level only result in emissions being moved within the ETS without affecting total emissions.

5.1.1 Effects of CO2 Taxation It is not easy from relatively limited experience to draw firm conclusions about effects – such a task really requires solid research. Typically, the research that does exist has been carried out for homogeneous industries9 rather than for the much more messy and diverse experience of a whole nation. Where research has been undertaken on the macro level of the economy, it is generally shown that carbon taxes do lead to reduced emissions (see Lin and Li, 2011; or Somanathan et al., 2014). As a simple illustration, Fig. 1 shows how the Swedish economy decarbonized much faster than the US, OECD or World averages in the period 1970–2014 in spite of starting at an unusually low level. Other countries with carbon taxes (although not as high as the Swedish taxes, such as the UK, Norway, and Denmark) also tend to have lower emission levels. 9 In Section 5.2 we review research from the transport fuel sector where there is more experience.

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FIGURE 1 CO2 emissions per GDP (Kg CO2 emissions per 2010 US dollars of GDP). Source: Own calculations based on emissions and GDP data from the World Development Indicators.

There is no strong theoretical reason to expect that carbon taxes or similar instruments would have effects on unemployment, deficits, growth or other macro variables. Nor would such effects be easy to identify and separate from other concurrent events in the economy. However, simplistic notions that carbon taxes are very damaging are quite common but they have little support since the economies like Sweden with high carbon taxes are in fact not performing worse, but if anything better, than other European countries. Turning to the microeconomic level of household-level and firm-level decisions, the fall in carbon intensity is particularly noticeable in the sectors where the full tax is applied. This is the case for example in the residential and commercial sectors, where Sweden, in spite of (or perhaps partly because of) a harsh climate and high energy taxation, uses no more energy for heating than countries farther south in Europe. The development is particularly noticeable in the district heating sector. District heating is itself an efficient way to heat houses because it allows for centralized heat production and further this technology can make use of waste heat, co-generation, heat pumps and other efficient technology. Sweden has a long tradition of district heating, which has expanded very significantly (roughly fourfold since 1970). Currently over 90% of all Swedish apartments are heated this way. The corresponding figure for space heating in the service sector is over 80%. At the same time as the system has expanded, it has also substituted its primary fuels, switching from fossil fuels to biofuels (e.g., wood residues and pellets) and household waste. Biofuels also have a large market share for heating in those houses that are not connected to district heating grids. The transition from fossil fuels to district heating and biofuels has been facilitated by targeted grants. However, the foremost factor responsible for the transition has been increased cost of fossil fuels due to CO2 taxes (Naturvårdsverket (Swedish EPA), 2004).

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In industry and other sectors, the use of biofuels and residues also has increased significantly, while the overall use of coal in the Swedish economy has been cut by more than 50% since its peak in the 1980s. Another sector where important long-run effects are noticeable is the transport sector, but here the effects are common to the whole of Europe; see Section 5.2.

5.2 TAXING (AND SUBSIDIZING) TRANSPORT FUEL This section discusses the effects of the taxation of transport fuels. Although general carbon taxes are unusual, a number of policies such as gasoline and diesel taxes have very similar effects. These taxes are applied to a subset of carbon fuels, yet they do cover a significant proportion of emissions in many countries. Fuel taxes are important for climate change mitigation because the transport sector represents a large and increasing share of carbon emissions (over 25% of global energy-related CO2 emissions in 2010; see Somanathan et al., 2014). In some countries, the fuel tax is clearly motivated by climate or other environmental factors. In others, it is not; the stated objective may instead be, for example, to finance road building. How do we decide when a tax is an environmental tax? To be frank, there is no bullet-proof definition. We will, however, take the approach here that fuel taxes can be considered environmental taxes, even though this may not have been their original stated purpose. The reason for this is that the effect of a tax will be largely independent of the motivation. A gas tax of US$1 will have the same effect on the climate whether motivated by climate, fiscal or roadbuilding reasons. We need to understand how a carbon tax would work if it could be enacted, and historical and geographic variation in fuel taxes actually does provide a perfect laboratory. For various reasons, different governments have at different times taxed gasoline and diesel at very different rates and researchers can use this variation to estimate the effects of a carbon tax. While the US federal, state, and local taxes average around US$0.50/gallon, many European countries have over seven times higher tax or around US$3.50/gallon (Parry and Small, 2005; Sterner, 2007). This amounts to an environmental tax of roughly US$400/ton of CO2 , implying that a large share of environmental revenues come from transport fuel taxes, which are common in Europe and Japan, as well as in lower-income, oil-importing countries. Fuel taxes thus seem easier to implement for the policymaker in some countries than other policies, partly because transport is not subject to much international competition and hence leakage rates are low. We can compare this situation with air travel, where planes can fuel in other countries and where politicians find that fuel taxation is much more difficult. Irrespective of their official motivation, the effect of taxes on fuel is clearly to raise prices to consumers and this is bound to reduce demand and thus act as if it were a climate-related carbon tax. Coincidental variation in policies in different countries, and occasional dramatic fluctuations in fuel prices, have provided a laboratory allowing us to gauge the effects of fuel prices (and hence taxes) on demand. Estimating

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fuel demand elasticities is quite a large industry and there are literally thousands of studies. There is of course a range of results depending on various factors. Fuel demand depends on income, urban architecture, population density, and the availability of alternative modes of transport. In the long run, these factors tend to influence each other, as higher fuel prices will tend to encourage people to live in denser cities and move closer to their jobs. They may also make public transport more profitable. There is also a good deal of inertia in demand patterns and hence elasticities vary significantly depending on the time horizon. Short-run elasticities may typically be as low as −0.1 to −0.2; a 10% increase in fuel price, for example, would decrease fuel consumption by only 1 or 2%, because people do not adjust their transportation patterns quickly. However, long-run elasticities tend to be in the range of −0.7. Thus, the high fuel taxes in Europe have probably been by far the most effective policy when it comes to actually reducing carbon emissions during the last three or four decades. Sterner (2007) estimates that if Europe had not followed a policy of high fuel taxation but had equally low taxes as the USA, then fuel demand would have been twice as large. Recently, new approaches in behavioral economics have focused on possible discrepancies between the impacts of taxes and the variation in fuel prices, due simply to variations in the oil market (Li et al., 2014). A recent study of the British Columbia carbon tax on transport CO2 emissions shows much bigger sensitivity (i.e. price elasticity) when the price of fuel is raised due to taxes (Rivers and Schaufele, 2015). Global variation in fuel prices is generated at least as much by subsidies as by taxes. Fossil fuel subsidies are still quite common in many countries, particularly in oil and coal producing areas. The IEA estimates the value of global fossil fuel consumption subsidies in 2016 around US$260 billion, with vast differences by country (IEA, 2016). Estimates vary widely by organization, depending on the methodology and whether externalities are included. (Note that some estimates of fossil fuel subsidies include tax exemptions, which is not always appropriate.) For example, the IMF estimated that energy subsidies globally totaled US$325 billion in 2015. A much broader and more general measure that includes unpaid externalities or damages was estimated at US$5.3 trillion, or 6.5% of global GDP (IMF, 2016). Ellis (2010) notes that subsidy removal would be akin to increasing global GDP by around 0.7% per year to 2050. The removal of these subsidies would also lead to a 15% reduction in global (energy related) CO2 emissions. One often mentioned problem of fuel taxation is its political economy. There is often considerable political resistance to fuel taxation and very often truckers and other special interests have been very active in lobbying against such taxes, making them difficult to implement. In this connection, it is often asserted that fuel taxes would be regressive. This is a good argument to mobilize opinion but not necessarily true, particularly not in the majority of countries that are low- or middle-income, where private transport is generally a luxury good, implying that it is higher-income people who spend more of their income on this good (Sterner, 2007). Very often, the poorest would be better off if the state raised revenue through taxing transport fuels than, for instance, value-added or some other broad-based tax. However, for the

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political feasibility of a tax, it is not really the welfare of the poor that matters most but the opinions of the most vociferous and influential groups, such as lobbies and the upper- and middle-income urban elites.

5.3 CAP AND TRADE SCHEMES In the last decade, more and more countries have instigated Emissions Trading Schemes (ETS) as important instruments to reduce emissions for climate-related issues and in some cases for other environmental or resource problems. The European Union touts its ETS as a mainstay of climate policy. In 2017, the World Bank estimated that there were 36 national level ETS programs implemented or scheduled for implementation (including those in the EU, Switzerland, Kazakhstan, and New Zealand) and 25 sub-national ETS programs (including in California, the US Regional Greenhouse Gas Initiative (RGGI) and Quebec) (World Bank and Ecofys, 2017). There is much debate among both economists and policymakers about the relative merits of taxes and ETS as instruments of climate policy. Clearly each has some advantages and some drawbacks. To start with the most pragmatic message: Any reasonably ambitious program will imply real resource costs and, if there is heterogeneity in abatements costs (which is the norm), then considerable savings can be achieved if an efficient market allocation is used – which implies that there must be a price on the pollutant, meaning carbon and other gases in the case of climate change. The choice of instrument, in this perspective, is secondary. The important thing is that we quickly attain a sufficiently high emissions price to fully internalize the externality and make real progress in dealing with climate change. Then, the choice becomes a pragmatic one: which instrument is politically most feasible. Experience to date shows that, if the programs are well designed, ETS can be effective, workable, and transparent tools for abatement in ways that mobilize the business sector, attract investment, and encourage international cooperation. At the same time, it is clear that both taxes and trading schemes face severe challenges in practice. Taxes are unpopular in general and resisted not only by the general public but more specifically by lobbies, as discussed in previous sections. In addition to this, it is hard to introduce taxes at the supranational level. Before creating the EU ETS, the European Union tried for several years to create a European-wide carbon tax. However, this was legally and politically very difficult. One might argue that this was a quarter century ago and that the gravity of the climate issue had not fully dawned upon all the member countries. But still today, the nation states are concerned because they zealously protect their taxation prerogative. Also, cap and trade faces a range of challenges, as we discuss below. This has induced at least some participants to think that it may actually on balance be easier to negotiate a global climate treaty that is focused on a joint (minimum) national carbon tax level than to negotiate one on quantities (Weitzman, 2017). In the textbook ETS, permits apply to all pollutants and all jurisdictions and for an unlimited period of time. Under such circumstances, one can demonstrate a lot

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of efficiency properties. However, in reality ETS tend to apply to some pollutants or some polluting activities but not others; they are typically decided for a limited “commitment period” and apply in a certain jurisdiction. Extending an ETS to cover more polluters or more pollutants is complex. Dealing with the changeover in commitment periods is also complex. The issue of banking (or even more so of borrowing) of credits is equally challenging, particularly between periods, though there have been cases of successful banking in other jurisdictions, e.g. the SO2 cap and trade program in the United States (Schmalensee and Stavins, 2013). Linking ETS schemes in different jurisdictions is similarly very complex because it opens up the Pandora’s box of the allocation of permits among countries – and thus indirectly of burden sharing between countries. Naturally, to maximize effectiveness, an ETS needs to be suitably designed in relation to its goals and context. Judging the success of an ETS program by the size of the price signal seems natural to many – to support this principle we can think of a tax of the same magnitude as the “dual” solution. In a simple programming sense, a tax of T should produce abatement roughly equivalent to a cap and trade scheme that has an equilibrium price of T. If the value T is judged unsatisfactory, then one could conclude that the ETS is not working well. Indeed, many observers in Europe are deeply disappointed by the very low EU ETS price signal of around US$5–10/ton CO2 . Recently reforms to the EU ETS, which imply retiring a large number of rights, have actually managed to raise the observed price. However, it did require a very considerable political fight. It is worth comparing with the highest carbon taxes mentioned earlier or the Obama administration estimate of the social cost of carbon, which was quite a bit higher at US$40 – but that has never been considered a feasible tax in the US and has not been reached in any mayor ETS either. If the tax was not feasible and the ETS is feasible, then clearly the ETS is better even if the price is low. The ETS also has a very different relationship to uncertainty. It gives certainty to a given emissions reduction but the agents of the economy have to live with uncertainty about the price signal. This has implications. When the price is too low, the environmentalists will complain. On the other hand it could be much worse if the prices rise too high – because then heavyweight economic interests may move in and disband the program altogether. The fear of this happening, together with the corruption and lobbying in the allocation process, tends to lead to an overallocation of permits with resulting low permit prices. This may be reinforced by pessimism concerning abatement possibilities. When the ETS is in place, firms often “discover” abatement options that were cheap and efficient which they did not know about. This is in fact one of the great advantages of a market mechanism. When these mechanisms combine with chance occurrences of sluggish growth, they may lead to a large excess supply of permits and a fall in the market. Other vicious circles can reinforce those mentioned. Decision makers become skeptical of the power of cap and trade to solve the problem and they start legislating or enacting additional programs of support for abatement, support for new technology, green certificates, mandates for energy efficiency and renewables, and so forth.

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All the additional programs, of course, may tend to lead to abatement – but within a cap and trade program, abatement in one firm or one sector just means that the price of permits is depressed and more emissions happen in another sector or country covered. All of this means that an ETS cannot, in fact, be judged simply by its permit price. The complications are compounded if, as mentioned, we have several gases, long time periods, banking between periods, and linking of permit schemes to each other. One can make a case that permit trading will work better if it applies to a larger geographic (and/or sectoral) coverage, but this is a simplification. Size can give stability but it can also easily propagate design errors. Lately there has been a number of interesting suggestions for improving the design of emission trading schemes. By giving the price of permits a floor and a ceiling, many of the disadvantages of uncertainty and variability can be combated. Theoretically, one can speak of a zone of hybrid instruments between a price-type and a quantity-type instrument. For example, there could be permit trading with quantity goals but price floors and ceilings which can be likened to elements of taxation and subsidization. Similarly, trading schemes can be linked to varying degrees (Burtraw et al., 2013; Green et al., 2014).

5.4 REFUNDING EMISSION PAYMENTS In theory, a uniform tax on all emissions is cost-effective if emissions are uniformly mixed. However Pigouvian taxes are not well understood nor liked: It may be that people resist the idea that their own behavior can be affected. They believe they only eat, drive, and consume what is necessary. For instance, they may say: “Look, my house is here, the daycare over there and my job on the other side of town. There is no public transport and so a fuel tax will not affect my miles driven but just be a transfer to the budget”. Econometric evidence shows that people do actually adjust behavior, but it seems that the argument is not popular (Klenert et al., 2017). For transboundary pollutants, we have problems of carbon leakage and often additional problems such as those created by cut-offs – the policymaker may want to impose a charge on some but not all firms. They may want to exempt small firms, or old firms, or firms in some regions or industries, etc. This creates additional problems of exit and entry of firms and control of pollution. The policymakers may also desire the “flexibility” afforded by giving out permits in a permit scheme to “pay off” political opposition. In all these cases, it is often overlooked that there are exactly analogous policies for price type instruments. For example, emission taxes can be made politically more feasible by exempting some firms from some proportion of the fees or taxes they have to pay and/or by earmarking the tax revenues and using them in some way that the polluting industries appreciate. Carbon pricing is becoming more popular, perhaps under the influence of the Paris Agenda. There are currently about 70 national or subnational schemes, with revenues of about US$26 billion in 2015 (World Bank et al., 2016; World Bank and Ecofys, 2017). However, concerns about competitiveness and carbon leakage stand in the way of wider adoption and higher carbon price levels (Aldy and Stavins, 2012;

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Aldy and Pizer, 2015; Ward et al., 2015). Many countries that do have carbon taxes have felt pressured to grant energy-intensive sectors exemptions or give them free permit allocations (Martin et al., 2014). Refunding the tax revenue can significantly ease the burden on the polluting firms and thereby reduce political resistance to the tax. Refunding emissions implies turning a tax into a fee, and then rebating it to polluters in some way. Tax revenues typically go into the national budget. Revenues from a fee can be refunded or be seen as payment for a service performed by a government agency or municipality. They can also be refunded directly to taxpayers. Kallbekken et al. (2011) show that citizens often fail to understand the Pigouvian aspect of carbon prices and only their revenueraising effect. When carbon revenues go into the general budget, some studies find public acceptance lower (Baranzini and Carattini, 2017). Sweden, with its generally accepted and very high carbon tax, may in this respect be an outlier that shows the importance of the process of fiscal reform by which the carbon tax is implemented (see Section 5.1). The clearest example of refunding is output-based refunding (OB, refunding in proportion to output) analogous to output-based allocation of permits (as was the case with the lead phase-out program in the USA, Tietenberg, 2003). Sweden has pioneered the use of OB for nitrogen oxides (NOx ). NOx is a byproduct of any combustion, whether oil, gas, coal or even biofuel, and in turn the NOx leads to acid rain, eutrophication, and other environmental problems. Just like sulfur oxides, this leads to health effects, and most industrialized countries have tried to reduce emissions. In Europe, this is regulated by the Gothenburg Protocol. Among the reasons for the use of refunded emission payments (REP) rather than a tax was the difficulty in dealing with a strong industrial lobby which would have opposed the tax virulently, arguing that companies would move abroad. It was thus impossible to set a sufficiently high tax on emissions to motivate abatement (Sterner and Isaksson, 2006). There is quite a large literature on the economics of output allocation; see, e.g., Fischer (2001), Fischer and Fox (2007), Gersbach and Requate (2004). Indeed, output-based refunding corresponds to output allocation or “benchmarking” in a tradable permit system (Fischer, 2001). OB allocation or refunding is often criticized because it generates an output subsidy and thus gives incentives for excess production. This effect is harmful in a competitive environment, but it can increase welfare under imperfect competition, where output is already suboptimally low due to restricted competition (Gersbach and Requate, 2004). Benchmarking in unilateral CO2 emissions policies is motivated by its potential to reduce carbon leakage and loss of competitiveness (Fischer and Fox, 2007). Refunding emission payments is often the only way of making a high emission fee politically feasible: It changes the political economy of policy making; the lobbies that form in opposition to a tax are very much weakened when faced by a REP because roughly half the firms in an industry actually make money from the scheme and thus tend not to oppose it (Fredriksson and Sterner, 2005). This is, in fact, the mechanism that allows the regulator to legislate a fee that is sufficiently high to have a real environmental effect.

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There are some alternative ways of buying support or reducing resistance that do not imply actual refunding of all the fees in proportion to output. In the French taxe parafiscale, the fees for NOx , SO2 , HCl, and VOCs were used generally to subsidize research and abatement (see Millock et al., 2004) in a more general sense. In the Norwegian system, a private NOx fund was created; it is allowed to collect the fees and use them to subsidize actual abatement costs at the firm level. This could be called a tax-subsidy combination or expenditure-based (EB) refunding. It can be shown that, if the OB system makes a high fee politically possible, the EB refunding makes it possible to reach the same environmental goal with a low fee. The efficiency of the low fee is heavily enhanced by the fact that the collected funds are used to subsidize abatement. Thus, this is a combination of a small fee and a subsidy for abatement equipment; see Hagem et al. (2012). Turning back to the bigger climate issues, revenue refunding straight to the polluters in proportion to output would probably not be practical because there are very many large and small polluters and it would be hard to define “output”. In this context, general revenue recycling is more reasonable. This is of course exactly what happens if the tax is put in the general budget and/or replaces other general taxes. Behavioral research has shown that the earmarking or refunding has to be salient and clear. It is important to label the tax a “fee” or a “climate contribution” and, to enhance acceptability, it is important to make the revenue recycling highly visible. The most obvious way may be through frequent and equal per capita transfers – perhaps in the form of checks; see further Klenert et al. (2017), Kallbekken et al. (2011), and Nature Editorial (2017).

5.5 REGULATION VERSUS TAXATION: THE EXAMPLE OF A HAZARDOUS CHEMICAL While environmental economists tend to think of cap and trade versus Pigouvian tax, the bulk of actual environmental regulation is focused on zoning, banning, phasing out, setting safe minimum standards, and other regulatory principles. Sometimes different countries choose different approaches and this gives us some opportunity to learn about the relative effectiveness of regulation versus price-based policies. One such case is the solvent trichloroethylene (TCE). Solvents are important industrial chemicals but can cause severe environmental and health problems because they are hazardous or toxic. One example is the effect of CFCs on the ozone layer. Other chemicals are persistent and bio-accumulating chemicals such as DDT (dichlorodiphenyltrichloroethane) or PCBs (polychlorinated biphenyls). Policymaking is difficult because there is genuine uncertainty about properties, costs, and alternatives. This section compares policy responses in some European countries to TCE and similar solvents which were banned in Sweden, taxed in Norway and Denmark, and strictly regulated in Germany. There are many possible policy instruments for policymakers to choose from: from taxes, charges, and deposit refunds, to tradable permit schemes, information provision, eco-labeling, liability legislation, refunded emissions payments, subsidies,

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and voluntary agreements (see Section 3). If there is a risk of serious and irreversible damages, the precautionary principle might seem to point in the direction of radical instruments like a prohibition. But if prohibition is not effective and leads to lobbying or “cheating” rather than research into new technologies, then market-based instruments that encourage such research may be more dynamically efficient. The Swedish parliament passed a law in 1991 banning the use of TCE in consumer products, starting in 1993, and prohibiting the professional use of TCE and methylene chloride, effective January 1, 1996. The Swedish case seems to illustrate the problems with such a strong instrument, which Sweden was alone in instituting: the ban is so absolute that it creates strong opposition among some users, who either find it particularly difficult to replace TCE or simply disapprove of the timing or policy method. The Swedish experience has shown that some firms spent a great deal of effort appealing and lobbying against the ban in the media and the courts. An alternative instrument would be that adopted by Germany: a very strict standard for technical and workplace ambient air that encouraged and in principle required completely closed systems for operation and even storage and transport. Detailed statistics show that the use of TCE in Sweden had already fallen from about 9000 tons per year to 3000 tons by the time of the ban. This rapid reduction appears to have been due to an arsenal of strong policies pursued by the Swedish authorities, and shows the power of day to day regulation by local authorities – at least when the latter are very motivated. Yet, when the ban was implemented, TCE was not completely phased out. Rather, use fell quite slowly after 1992. Interviews showed that companies did not believe the authorities would succeed with a ban and so decided to fight it. A group of industries even published an open letter to the Prime Minister in a leading Swedish newspaper, asking for a repeal of the ban and threatening to move abroad if the ban were enforced. In the end, many companies were given waivers, often after appealing to EU courts, and this explains why use did not fall rapidly. Detailed studies showed that most companies could easily and cheaply find alternatives to TCE. Fig. 2 shows reported marginal costs of abatement for a sample of companies and it is clear that most companies would have had an incentive to voluntarily stop using TCE if there were a tax or fee of SEK 50 per kilogram. Interestingly Norway did, later, introduce a tax per kilo on both TCE and PER of 50 Norwegian Krone, roughly the same as SEK 50. Denmark also has an environmental tax, though it is very low, and Germany has its tough regulation. Fig. 3 plots the rates of phase-out in the four countries compared with the rest of Europe. Although correlation is not proof of causation, all of the countries with stricter policies appear to have been quite effective compared to the majority of other European countries, whose policies were less stringent and where emissions have declined only very gradually. The conclusion appears to be that if the policymaker wants to avoid this health hazard then some policy instrument is needed, and that a ban is not necessarily more effective than economic instruments in quickly reducing emissions, because

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FIGURE 2 Marginal abatement cost and effects of tax compared with ban. Assumptions: 15-year equipment life, 4% real interest. Source: Slunge and Sterner (2001).

FIGURE 3 Rates of reduction of TCE. Source: Slunge and Sterner (2001).

it is tricky in practice to implement a ban effectively. A tax is easiest to administer, whilst the tough German regulations appear to have led to considerable technical development and even export opportunities for firms in Germany that manufacture particularly high-tech clean equipment. Ultimately, countries are likely to be influenced in their choice of policy according to their own legal and industrial traditions; see Sterner and Coria (2012, Ch. 20).

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5.6 POLICIES TO MODIFY BEHAVIORAL NORMS Consumer behavior is often driven by norms, habits, and culture. Where such behavior is detrimental to the environment (locally or globally), there is a role for public policy in changing cultural and behavioral norms. However, choosing the right policy mechanism is likely to be country-specific. Meat and tobacco are examples of commodities which are harmful to individual health when consumed/used in excess, and which are associated with negative environmental externalities. They are also products where use is strongly linked to culture and societal norms. Lessons can be learned from how the UK government changed cultural norms over smoking. In many countries, including the UK, smoking used to be prevalent in private and public spaces, including offices, pubs, and restaurants. In the 1970s, over half the adult male population and 40% of the adult female population in the UK smoked. This number fell to around one-third of the adult population in 1981 (Cairney, 2007) and to 16% in 2016 (UK Office for National Statistics). The UK government has relied on a number of policy instruments to reduce tobacco consumption; these span the four categories of pricing, rights, regulation, and information. Links between tobacco smoking and lung cancer were first rigorously demonstrated in the 1950s (Doll and Hill, 1956). The UK government started regulating tobacco in the early 20th century with a ban on the sale of tobacco to children under 16, and later a ban on TV advertising in 1965 (Cairney, 2007). Some regulations from the 1970s such as “non-smoking areas” in airplanes seem quite laughable today. Effective bans have followed more recently, such as a smoking ban in public places, including pubs, fully implemented in the UK in 2007. Note that this was motivated to a large extent as an environmental issue: to protect people from secondhand smoke. This ban was opposed by some. But the reality is that in 2007 there was already considerable support for the ban, as reported in the UK newspapers. The UK’s Office of National Statistics found that 77% of people agreed with the new legislation, and only 15% disagreed. Whereas 8% of those surveyed said that they would visit pubs less often, 15% said that they would visit more. These data suggest that the new legislation was easier to implement because public opinion already supported the proposed changes, and indeed many smokers wanted to quit and so welcomed legislation that would make continuing to smoke more difficult and less socially accepted. Governments have also long taxed tobacco. A tax on an addictive substance arguably acts less as a Pigouvian tax and more as a source of government revenue. Tax currently accounts for around 89% of the cost of a packet of cigarettes but price elasticities are low and effects small. Some have argued this might make governments less interested in really reducing the habits, though there is little evidence for this. The UK government increasingly is also using a number of nudges, which can “help to promote a culture that is accepting of legislation to promote health” (Marteau et al., 2011). For example, advertising campaigns have cast non-smoking and wanting to give up smoking as the norm; cigarettes are kept out of sight in retail outlets; cigarette packages now contain health warnings, often paired with pictures of diseased lungs; and standardized packaging was introduced in 2016.

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As norms have changed over time, and the location of consumption has changed as well, the relative importance of each policy instrument has changed, reflecting the shift of tobacco use away from the public sphere and an overall drop in consumption, with some conversion to e-cigarettes. Taxation has been progressively increased over time; the number of locations where smoking is permitted has fallen; and smoking has become less socially acceptable. Thus, public policy with regard to smoking has shifted from a focus on controls and increased prices, to a broader approach that also incorporates behavioral nudges. Similarly, social policies that change the social acceptability of smoking have been shown to be as important as taxation in reducing smoking rates in the US (Alamar and Glantz, 2006). Taxation has particular implications for equity when the price elasticity of demand is low; when the good being taxed is addictive, as is the case for tobacco; and when use is more concentrated among lower-income individuals. Ultimately, whether such taxes are progressive or regressive is an empirical question, and depends on the budget shares and relative price sensitivity of higher and lower income groups (Colman and Remler, 2008). It also depends on the incidence of benefits. Meat and dairy are food products that have particularly high greenhouse gas footprints, in addition to contributing to water scarcity and erosion. Reducing meat consumption would have both climate and health benefits (Watts et al., 2018). China is aiming to reduce meat consumption by 50% by 2030, so as to reduce greenhouse gas emissions and the increasing incidence of obesity and diabetes in the country (Milman and Leavenworth, 2016). Yet meat consumption is linked to socio-economic status and may have strong cultural links. For example, consumption is typically found to increase in countries as incomes increase and populations become more urbanized (Fiala, 2008). The lessons from the UK’s efforts to reduce tobacco consumption suggest a combination of taxation and societal nudges may be required. There is some evidence that, though reducing meat consumption is an important climate change mitigation strategy, governments in higher-income countries have not yet acted, and efforts led by NGOs are linked to health benefits rather than climate (Laestadius et al., 2014). Given the increasing evidence concerning the relative health and social impacts of various stimuli (alcohol, tobacco, narcotic drugs), the different policy responses to each in many higher-income countries often reflect historical biases and tones of morality rather than rational policy making. Whichever policies are chosen by government should, at a minimum, recognize explicitly the costs, benefits, harms, and unintended consequences of those policies. Domestic energy use choices similarly have important implications for the environment, yet consumers have been found to be “slow to habituate” to adopting more environmentally friendly choices (Allcott and Rogers, 2014, p. 3003). In such a circumstance, one role of government policy is directly to help individuals to change their habits, or to encourage utility companies to do the same. Presenting individuals with social comparisons is one approach has been demonstrated to change behavior across a wide variety of situations, including with respect to reducing energy use (Ayres et al., 2013).

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6 DESIGNING POLICIES FOR THE ANTHROPOCENE Research in Earth systems science shows that we face numerous large-scale environmental problems. Today humankind constitutes the largest driver of change at the planetary scale and the implications for global ecosystem stability and for society are profound. Social impacts often hit far away from the source of the problem and we risk crossing tipping points, with potentially catastrophic costs. A number of Geoscientists and Earth system scientists speak of the Anthropocene – the era when human activities and the global economy are the dominant forces of change that threaten the stability and resilience of the entire Earth system. Steffen et al. (2015) enumerate nine planetary boundaries that interact to regulate the overall stability of the planet. One of these boundaries, the composition of the atmosphere leading to climate change, is already proving to be a very hard problem for the nations of the world to deal with. Progress over the past 25 years has been insufficient. Similarly, biodiversity loss, ocean acidification, atmospheric aerosol loading, and stratospheric ozone depletion, are all pushing the Earth system up against planetary boundaries. There is controversy and debate concerning this work but no doubt that global ecosystems are under stress. The Millennium Ecosystem Assessment (2005) states: “Over the past 50 years, humans have changed ecosystems more rapidly and extensively than in any comparable period of time in human history. This has resulted in a substantial and largely irreversible loss in the diversity of life on Earth. The degradation of ecosystem services could grow significantly worse during the first half of this century and is a barrier to achieving the Millennium Development Goals. . . . The challenge of reversing the degradation of ecosystems while meeting increasing demands for their services can be partially met under some scenarios that the MA has considered, but these involve significant changes in policies, institutions, and practices that are not currently under way.” (Emphasis added by authors.)

The biosphere’s capacity for adaptation is being stretched to its limits, creating potential risks of sudden collapse. To stay within planetary boundaries, we must understand why we have a tendency to transgress boundaries. The fact that the human economy has pushed us up against, and in some cases perhaps past, some of the boundaries that define a safe operating space means that very skillful maneuvering will be required (Rockström et al., 2009). Economists, social, behavioral, and policy scientists must expand our analysis significantly to face the drastic changes needed in how we manage the ecosystem resources of this planet. Environmental policy design and related fields typically deal with one manageable problem within one single (national or maybe federal) jurisdiction. When dealing with planetary problems, it is seldom possible to implement traditional policy instruments, regulations or taxes, although there are large potential gains of taxing “societal ills” to generate revenues to meet environmental challenges. Serious concern about planetary-scale problems is rather recent in natural science and virtually non-existent in economic models, most of which disregard boundaries, ecological

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tipping points, and other complexities or uncertainties. Broadening the analysis of environmental policy making to deal systematically with policies to keep the Earth system within planetary boundaries is challenging because we need to deal simultaneously with several difficult aspects, which implies going far beyond the state of the art. We are speaking of an extreme degree of multi-pollutant, multi-governance, and multi-policy setting that will necessitate new, high-risk extensions in several different dimensions that are the major challenges we need to address: A. B. C. D.

Expansion of geographic and political scope Significant extension in time scale Significant extension in number of pollutants and scientific complexity Proper inclusion of equity, ethics, risk, uncertainty, and governance concerns

6.1 AN EXPANSION OF GEOGRAPHIC AND POLITICAL SCOPE Dealing with planetary boundaries requires a global response. This is not a simple matter of “scaling up” national instruments or applying the lessons from the federalism literature. Many of the most pressing environmental challenges imply an urgent need for collective action among multiple actors. For problems that are truly global, such as the climate issue, the stratospheric ozone issue and – by their very nature – most of the planetary boundary issues, most effective policies have to be formulated at a global level. Yet because there is very little power at global or international levels due to state sovereignty, most international policy making is obliged to take the form of international agreements, negotiations, and treaties among countries, concerning at least goals, and possibly also the application of policy instruments at the national level. Important elements in finding solutions for such critical transboundary problems that have been identified include dialogue that includes scientists and officials; “complex, redundant, and layered institutions”; and flexibility with regard to mixing institutional types, learning, experimentation, and change (Dietz et al., 2003). There are already several examples of long-standing international treaties created to manage global commons, natural resources shared by the globe as a whole, comprising the high seas, the atmosphere, Antarctica, and outer space (Ostrom et al., 1999; Buck, 2017). Two noteworthy examples cover the Antarctic (The Antarctic Treaty, see also Vogler, 2012) and Outer Space (The Outer Space Treaty of 1967; see also Chaddha, 2010), yet these have only required the cooperation of a small number of countries. More recently the Montreal Protocol on stratospheric ozone, signed in 1987, has proven a successful international governance regime (Dietz et al., 2003). For other global commons such as the climate, biodiversity, forests, and the oceans, global participation in any regulation is typically required, because most if not all countries benefit from and/or degrade the resources. Biodiversity has been addressed on many scales and by a variety of policy instruments (see evaluations done by e.g. Wätzold and Schwerdtner, 2005; Miteva et al., 2012). The Convention on Biodiversity, adopted by most countries in 1992, follows the model of employing international conventions for global transboundary issues.

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Global climate change has attracted significant analysis (see IPCC, 2014), though there is insufficient action in the political sphere. The United Nations Framework Convention on Climate Change (UNFCCC) Kyoto Protocol codifies the international climate change policy regime, sets targets for Annex 1 (industrialized) countries to reduce their GHG emissions, and allows for trading carbon credits in international markets. Lower-income countries can participate in these markets through the supply of carbon credits (certified emission reductions) through CDM (the Clean Development Mechanism) and more recently, through REDD (Reduced Emissions from Deforestation and forest Degradation). Each of these initiatives rewards lowerincome countries for reducing GHG emissions, through the transfer of funds from higher-income countries. In reality, the markets created by these policy mechanisms have not quite lived up to the original optimistic expectations. Theory generally has quite a pessimistic view of the possibility of reaching successful and sufficiently radical international environmental agreements due to national sovereignty: nations are always free to leave a treaty. Therefore, treaties have to be designed so that nations find it in their interest to remain and comply, which naturally puts limits on how far a treaty can be constructed to constrain behavior or emissions that are damaging to ecosystems. There is a considerable literature concerning how to get around this point by clever design of the treaty or by linking to other issues such as trade (see Barrett, 2003). Barrett (2013) has specifically analyzed the importance of discontinuous benefits, as would be given by the existence of an ecological threshold. He finds that if the threshold is known with certainty and the damage sufficiently large, then countries will coordinate to avoid catastrophe. Under these circumstances, climate treaties can sustain the efficient outcome. If, however, there is uncertainty – even moderate uncertainty – then the results can break down (Barrett and Dannenberg, 2014). In the implementation of international treaties, it is vital that all (or practically all) countries participate – not only for fairness, but because we face the problem of leakage. A small country with, say, 10% of world production, acting alone, can never hope to have more than a 10% effect on a global problem. Moreover, with the textbook example of leakage (e.g. carbon leakage), the effect may be much smaller than that because the production, jobs, and pollution will just increase in another country (see e.g. Hoel, 1991). More generally, spatial and temporal aspects, such as equity considerations across higher- and lower-income countries, and across current and future generations, are all very hard to deal with. For example, rapidly industrializing countries such as China and India, which currently have relatively low per-capita emissions, though high CO2 emissions in absolute terms, did not have obligations for emissions reductions in the 1997 Kyoto Protocol (Pittel and Rübbelke, 2008). However, the 2015 Paris agreement required commitments from all the 195 signatories, which include both higher and lower-income countries. Issues of measurement, accountability, regulation, and liability similarly come to the fore. For global commons, the reality of individual incentives differing from group incentives tends to be considerably amplified.

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Successfully mitigating the adverse effects of free-riding on a large scale is another difficulty in international policy-making. Nordhaus (2015) suggests a solution which he refers to as ‘Climate Clubs’. The basic principle for a club is based on the notion that, without enforcement and the possibility to sanction defectors, international agreements such as the Kyoto Protocol will break down. Within a club, countries would have ambitious emission reductions targets (high carbon prices). Countries that would not join the club would, in contrast, have low abatement costs (low or zero carbon prices) but would then face penalties by the countries that are members of the club, for example through trade tariffs. The probability of sustaining a solution based on a ‘club regime’ hinges upon a modest carbon price (too high a price would make countries prefer the penalties); on the other hand, Nordhaus (2015) shows that only small trade penalties are necessary for significant abatement levels. As the excruciatingly slow progress of climate policy clearly shows, the making of an international treaty is no simple task, largely because there is no world government, no global policy making arena. There is no clear set of universal rights (for instance, rights to emit climate gases into the atmosphere – or rights to clean air or rights to a stable climate regime). A direct consequence is that the primary set of instruments is radically different from that used in national policy making. We need to develop our understanding of treaty negotiations, remembering that these treaties, in turn, may require the use of national instruments. They must certainly coexist with the national instruments that countries choose as vehicles to implement the treaties and therefore we need to consider the linkage compatibility of instruments with various types of treaties (Stavins, 2016). Other instruments that focus on cultural and existential drivers, on the scale of the economy, on behavior, information, finance, culture, population demographics, and technology, may be important to truly transform rather than just modify societies (Barbier, 2015). Innovation – technical, organizational, and cultural – is similarly important. It may seem contradictory, but actions to stay within planetary boundaries will not only require action at the global scale, but also locally. Take as an example the disturbances to bio-geophysical flows of phosphorous or nitrogen, or land management decisions in farming or forestry with effects on emissions of climate forcing gases such as methane or nitrous oxides. Most of these emissions stem from non-point sources – emissions that cannot be easily measured or verified by national authorities. Even though the latter have formal jurisdiction over the emissions, the most realistic approach may be local Common Property Resource management (Ostrom, 1990). It is not uncommon that the same activity (say, forestry) can have effects on both local and global pollution and would thus possibly be subject to regulation at several levels simultaneously, which can lead to a fatal lack of clarity in rulemaking. Also, ecosystem health can depend on changes at a number of different geographical (and time) scales. The health of a coral reef depends not only on global climate and acidification but on local processes of erosion, pollution or overfishing. This is an area where we may be able to build on Ostrom’s theories on polycentric decision-making (Ostrom, 2009).

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By definition, policy instruments for managing global commons will almost certainly involve in some way lower-income countries. However, these countries will also need to design and implement environmental policy instruments to deal with national and sub-national environmental issues. It is crucial to consider the lowerincome country (“developing world”) perspective, because lower-income countries generally are characterized not only by inequality and poverty, but also by weak institutions, poor performance of markets, lack of data – in particular environmental data, and typically, lack of experience on design and implementation of environmental policies (Sterner and Somanathan, 2006). In the 1990s, Afsah et al. (1996) suggested that in many lower-income countries the pre-conditions for applying many policy instruments, including market-based instruments, did not yet exist. Regulators often did not have sufficient information, public mandate, nor capacity for priority setting. In many countries, this reality still exists. Further, whether public policy focuses on taxes, quantity restrictions, or regulation, governments must be able to enforce their policies, and this is particularly tricky in lower-income countries, where government departments tend to be underfunded, property rights tend to be poorly defined, and corruption and non-compliance may be a societal norm.

6.2 SIGNIFICANT EXTENSION IN TIME-SCALE Dealing with the environmental problems that threaten planetary boundaries implies significantly extending the time-frame of our analysis. A very long-run view is unusual in economics but imperative for this set of problems. The reason for this is the inertia of many planetary processes, such as the warming of the oceans. Longer time scales introduce numerous problems in policy making, including that of time consistency and commitment, in some ways analogous to including multiple jurisdictions. Just as there is no world government to force nations to comply, there is practically no mechanism by which current decision makers can commit future politicians to any particular course of action. Thus, a global treaty can stipulate carbon taxes that rise over time but nothing actually stops future governments from changing – or even leaving – the program. Creating property rights can, in some contexts, be a mechanism for creating long-run policy commitment and hence rights-based management might be seen as a shortcut to stability, but this feature makes them difficult to design – particularly in countries without a strong tradition of well-defined and enforced property rights. Longer time horizons not only raise the issues of time consistency and the “sovereignty” of each generation of politicians, they also raise issues of optimal resource use over time, as well as ethically contentious and risky issues, such as that of regulating economic growth and discussing population numbers. If the climate debate has succeeded in anything this far, one could point to a lengthening of our time perspective. It is now standard to consider developments up to the year 2100 – which is a lot longer than economic analyses have typically dealt with in the past. Still, hardly anybody is willing to seriously consider several centuries, which is clearly necessary when considering, for instance, climate-induced sea level

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rise. Conventional linear discounting at a rate of 3–5% diminishes all future values to virtual insignificance and thus is not appropriate in an economy that will be constrained from growing and where we must even face the risk that people may be impoverished. Ultimately, we need a vision of the future to set relevant prices, including social discount rates, that can vary over time and vary between sectors and instruments to guide society onto a sustainable path (Hoel and Sterner, 2007; Arrow et al., 2013).

6.3 SIGNIFICANT EXTENSION OF THE NUMBER OF POLLUTANTS AND SCIENTIFIC COMPLEXITY One needs also to consider that there is interaction between the drivers that are leading to transgression of the various planetary boundaries and local management of local ecosystems. For instance, the effect of, say, climate change on ecosystems such as coral reefs may depend on local actions, such as reductions in pollution runoff and fishing. This creates links between policy at the global and local levels and between different areas of concern or different pollutants/drivers. Thus we have multi-pollutant/multinational and multiple agents. One can in some cases see promise of good news: the incentives for countries to intervene in specific ways are stronger than the incentives to address the root cause of the problem. An example is perhaps given by those who argue that incentives to act on climate change by reducing coal use should be strengthened by the “ancillary benefits” of combating local healthissues and pollution related to coal (Watts et al., 2015, 2018). On the other hand, the level of complexity rises and the already considerable complexity of the planetary boundaries approach is further complicated. Policy making in the Anthropocene will mean working with multiple planetary boundaries as well as with the unintended side effects of other policies, for instance, to promote industry or agriculture. Many environmental issues will require close collaboration between natural and social scientists. There is some incipient work on the challenge of dealing with multiple pollutants (see Ambec and Coria, 2013). Generalizing this and extending it to cover the multiple challenges mentioned is a big step. There is clearly a risk involved in trying to take steps that are too big. The progress of science is typically by studying small extensions from given theory in order to stand on firm ground and to be able to prove analytically new results. On the other hand, it is sometimes possible to take a quantum leap and find completely novel ways of dealing with multiple problems.

6.4 EQUITY, ETHICS, RISK, UNCERTAINTY, AND GOVERNANCE Policy making for major environmental issues must deal with considerable uncertainty and risk. More risk, not only concerning the economy but more fundamentally ecological thresholds, will raise countless ethical and distributional issues, just as a longer time horizon does. Moreover, the poor are typically more vulnerable to uncer-

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tainties because they have fewer reserves and thus risks are tied to ethical and equity issues. At a global scale, the poor do appear to bear disproportionately the burden of many environmental problems – whether climate change, biodiversity or freshwater availability. Dealing with major global issues rather than standard domestic environmental issues will tend to exacerbate a whole series of problems related to the distribution of income and wealth. Indeed, policies to respect planetary boundaries will almost certainly have distributional implications within and among contemporary societies. When poor and very unequal societies are affected, issues of welfare and distribution will be more prominent, and perceived equity and justice issues dominant (Johansson-Stenman and Konow, 2010). The cultural settings of different nations are very varied and have profound implications for the acceptability of different policies (Barbier, 2011). Traditional policy instrument design tends to focus on issues of efficiency, but dealing with policy making for the Anthropocene forces us to focus on issues of distribution and fairness, which are decisive for political feasibility and for welfare consequences. Policy designed to deal with global environmental issues should allow scope for economic development of lower-income economies and must be compatible with poverty alleviation if it is to be globally acceptable. A common but misleading assumption in much of the policy literature is that we can calculate the effect of policy instruments by conventional neoclassical methods. In its most narrow sense, neoclassical economics assumes that human rationale is egoistic, individualistic and material. Hence, we can calculate profit or utility functions for firms or individuals and design instruments from the assumption that individuals are maximizing individual utility. However, we know this to be wrong: fairness and social norms are very important (Johansson-Stenman and Brekke, 2008). This brings us full circle to the start of this paper where we discussed the fundamental reasons for market failure and the ways in which economics describes human behavior. There are many areas where a more realistic description of human action than simple utility maximization should be used. Crucial issues include long-run choices over time under uncertainty and risk. These are typical for climate change and many other large environmental issues with which we are concerned and they are cases where behavioral economics has shown that conventional theory tends to do a poor job of analyzing human behavior. We are at a crossroads when the formulation of wise policy to deal with these global, complex, long-run planetary problems has become very challenging. The consequences for welfare can be very sizeable and very significant, making issues of distribution and fairness stark issues of survival. At this time, we need robust politics and strong systems of governance. Unfortunately there seem to be signs of the opposite. As the stakes get higher, petty conflicts over distribution escalate and politics runs the risk of degenerating into populism and strife.

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REFERENCES Afsah, S., Laplante, B., Wheeler, D., 1996. Controlling Industrial Pollution: A New Paradigm. World Bank Policy Research Working Paper No. 1672. 22 pp. Agrawal, A., 2007. Forests, governance, and sustainability: common property theory and its contributions. International Journal of the Commons 1 (1), 111–136. Alamar, B., Glantz, S.A., 2006. Effect of increased social unacceptability of cigarette smoking on reduction in cigarette consumption. American Journal of Public Health 96 (8), 1359–1363. Albers, H.J., 2010. Spatial modeling of extraction and enforcement in developing country protected areas. Resource and Energy Economics 32 (2), 165–179. Albers, H.J., Maloney, M., Robinson, E.J.Z., 2017. Economics in systematic conservation planning for lower-income countries: a literature review and assessment. International Review of Environmental and Resource Economics 10 (2), 145–182. Allcott, H., Rogers, T., 2014. The short-run and long-run effects of behavioral interventions: experimental evidence from energy conservation. The American Economic Review 104 (10), 3003–3037. Aldy, J.E., Pizer, W.A., 2015. The competitiveness impacts of climate change mitigation policies. Journal of the Association of Environmental and Resource Economists 2 (4), 565–595. Aldy, J.E., Stavins, R.N., 2012. The promise and problems of pricing carbon: theory and experience. The Journal of Environment & Development 21, 152–180. Aldy, J.E., Viscusi, W.K., 2014. Environmental risk and uncertainty. In: Handbook of the Economics of Risk and Uncertainty, vol. 1. North-Holland, pp. 601–649. Ambec, Stefan, Coria, Jessica, 2013. Prices vs quantities with multiple pollutants. Journal of Environmental Economics and Management 22 (1), 123–140. Antle, J., Capalbo, S., Mooney, S., Elliott, E., Paustian, K., 2003. Spatial heterogeneity, contract design, and the efficiency of carbon sequestration policies for agriculture. Journal of Environmental Economics and Management 46 (2), 231–250. Arrow, K., Cropper, M.L., Gollier, C., Groom, B., Heal, G.M., Newell, R.G., Nordhaus, W.D., Pindyck, R.S., Pizer, W.A., Portney, P., Sterner, T., Tol, R., Weitzman, M.L., 2013. Determining benefits and costs for future generations. Science 341 (6144), 349–350. Ayres, I., Raseman, S., Shih, A., 2013. Evidence from two large field experiments that peer comparison feedback can reduce residential energy usage. Journal of Law, Economics, & Organization 29 (5), 992–1022. Baranzini, A., Carattini, S., 2017. Effectiveness, earmarking and labeling: testing the acceptability of carbon taxes with survey data. Environmental Economics and Policy Studies 19, 197–227. Barbier, E.B., 2011. Transaction costs and the transition to environmentally sustainable development. Environmental Innovation and Societal Transitions 1 (1), 58–69. Barbier, E.B., 2015. Nature and Wealth: Overcoming Environmental Scarcity and Inequality. Palgrave MacMillan, London. Barrett, S., 1994. Strategic environmental policy and international trade. Journal of Public Economics 54 (3), 325–338. Barrett, S., 2003. Environment and Statecraft: The Strategy of Environmental Treaty-Making. OUP, Oxford. Barrett, S., 2013. Economic considerations for the eradication endgame. Philosophical Transactions of the Royal Society, Series B 368 (1623). https://doi.org/10.1098/rstb.2012.0149. Barrett, S., Dannenberg, A., 2014. Sensitivity of collective action to uncertainty about climate tipping points. Nature Climate Change 4, 36–39. https://doi.org/10.1038/nclimate2059. Bateman, I.J., et al., 2013. Bringing ecosystem services into economic decision making: land use in the United Kingdom. Science 341, 45–50. Baylis, K., Peplow, S., Rausser, G., Simon, L., 2008. Agri-environmental policies in the EU and United States: a comparison. Ecological Economics 65 (4), 753–764. Becker, G.S., 1983. A theory of competition among pressure groups for political influence. The Quarterly Journal of Economics 98 (3), 371–400. Berkes, F., 2006. From community-based resource management to complex systems. Ecology and Society 11 (1), 45. [Online] http://www.ecologyandsociety.org/vol11/iss1/art45/.

ARTICLE IN PRESS References

Bjertnæs, G.H., Fæhn, T., 2008. Energy taxation in a small, open economy: social efficiency gains versus industrial concerns. Energy Economics 30 (4), 2050–2071. Blackman, A., Osakwe, R., Alpizar, F., 2010. Fuel tax incidence in developing countries: the case of Costa Rica. Energy Policy 38 (5), 2208–2215. Bosetti, V., Carraro, C., Massetti, E., Tavoni, M., 2008. International energy R&D spillovers and the economics of greenhouse gas atmospheric stabilization. Energy Economics 30 (6), 2912–2929. Bovenberg, A.L., de Mooij, R.A., 1994a. Environmental levies and distortionary taxation. The American Economic Review 94 (4), 1085–1089. Bovenberg, A.L., de Mooij, R.A., 1994b. Environmental policy in a small open economy with distortionary labour taxes: a general equilibrium analysis. In: van Ierland, E.C. (Ed.), International Environmental Economics. Elsevier, Amsterdam. Bovenberg, A.L., de Mooij, R.A., 1997. Environmental tax reform and endogenous growth. Journal of Public Economics 63, 207–237. Bovenberg, A.L., Goulder, L.H., 1996. Optimal environmental taxation in the presence of other taxes: general equilibrium analyses. The American Economic Review 86 (4), 985–1000. Brock, W., Xepapadeas, A., 2010. Pattern formation, spatial externalities and regulation in coupled economic–ecological systems. Journal of Environmental Economics and Management 59, 149–164. Brown, G.M., 2000. Renewable natural resource management and use without markets. Journal of Economic Literature 38 (4), 875–914. Buck, S.J., 2017. The Global Commons: An Introduction. Routledge. Burtraw, D., Keyes, A., Zetterberg, L., 2018. Companion policies under capped systems and implications for efficiency – the North American experience and lessons in the EU context. Resources for the Future Report. June 2018. Available at: http://www.rff.org/files/document/file/ RFF-Rpt-Companion%20Policies%20and%20Carbon%20Pricing_0.pdf. Burtraw, D., Palmer, K.L., Munnings, C., Weber, P., Woerman, M., 2013. Linking by Degrees: Incremental Alignment of Cap-and-Trade Markets. Cabe, R., Herriges, J.A., 1992. The regulation of non-point-source pollution under imperfect and asymmetric information. Journal of Environmental Economics and Management 22 (2), 134–146. Cairney, P., 2007. A ‘multiple lenses’ approach to policy change: the case of tobacco policy in the UK. British Politics 2 (1), 45–68. Cárdenas, J.C., 2000. How do groups solve local commons dilemmas? Lessons from experimental economics in the field. Environment, Development and Sustainability 2 (3–4), 305–322. Cárdenas, J.C., 2016. Human behavior and the use of experiments to understand the agricultural, resource, and environmental challenges of the XXI century. Agricultural Economics 47 (S1), 61–71. Carlsson, F., Kataria, M., Krupnick, A., Lampi, E., Löfgren, Å., Qin, P., Sterner, T., 2013. A fair share: burden-sharing preferences in the United States and China. Resource and Energy Economics 35 (1), 1–17. Carpenter, S.R., Caraco, N.F., Correll, D.L., Howarth, R.W., Sharpley, A.N., Smith, V.H., 1998. Nonpoint pollution of surface waters with phosphorus and nitrogen. Ecological Applications 8 (3), 559–568. Carraro, C., Lévêque, F., 2013. Voluntary Approaches in Environmental Policy. Economics, Energy, and Environment. Springer Science and Business Media, Dordrecht. Cason, T.N., Gangadharan, L., 2013. Empowering neighbors versus imposing regulations: an experimental analysis of pollution reduction schemes. Journal of Environmental Economics and Management 65 (3), 469–484. Caswell, J.A., Mojduszka, E.M., 1996. Using informational labeling to influence the market for quality in food products. American Journal of Agricultural Economics 78 (5), 1248–1253. Chaddha, S., 2010. A tragedy of the space commons? Available at SSRN: https://ssrn.com/abstract= 1586643 or https://doi.org/10.2139/ssrn.1586643. Coase, R., 1960. The problem of social cost. The Journal of Law & Economics 3, 1–44. Cole, D., 2002. Pollution and Property: Comparing Ownership Institutions for Environmental Protection. Cambridge University Press, New York. Colman, G.J., Remler, D.K., 2008. Vertical equity consequences of very high cigarette tax increases: if the poor are the ones smoking, how could cigarette tax increases be progressive? Journal of Policy Analysis and Management 27 (2), 376–400.

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ARTICLE IN PRESS 48

Selection and design of environmental policy instruments

Comest, U., 2005. The Precautionary Principle. World Commission on the Ethics of Scientific Knowledge and Technology. Conniff, R., 2018. Why Green Groups Are Split on Subsidizing Carbon Capture Technology. Yale Environment 360, Published at the Yale School of Forestry & Environmental Studies, April 9. Conrad, J.M., 1999. The economics of nonrenewable resources. In: Resource Economics. Cambridge University Press, New York. Conrad, J.M., Clark, C.W., 1987. Natural Resource Economics. Cambridge University Press, New York. Coria, J., 2009. Taxes, permits, and the diffusion of a new technology. Resource and Energy Economics 31 (4), 249–271. Costello, C., Ovando, D., Clavelle, T., Strauss, C.K., Hilborn, R., Melnychuk, M.C., Branch, T.A., Gaines, S.D., Szuwalski, C.S., Cabral, R.B., Rader, D.N., Leland, A., 2016. Global fishery prospects under contrasting management regimes. Proceedings of the National Academy of Sciences of the United States of America 113 (18), 5125–5129. Cropper, M.L., Oates, W.E., 1992. Environmental economics: a survey. Journal of Economic Literature 30 (2), 675–740. Dasgupta, P., 1983. The Control of Resources. Oxford University Press. Dasgupta, P.S., Heal, G.M., 1979. Economic Theory and Exhaustible Resources. Cambridge Economic Handbooks. Demaria, F., Schneider, F., Sekulova, F., Martinez-Alier, J., 2013. What is degrowth? From an activist slogan to a social movement. Environmental Values 22 (2), 191–215. Derwort, P., 2016. If at first you don’t succeed. . . Institutional failure in the public sector. Sustainability Governance, 21st January 2016. https://sustainability-governance.net/tag/policy-failure/. Diamond, L., 2007. A quarter-century of promoting democracy. Journal of Democracy 18, 118–120. Dietz, T., Ostrom, E., Stern, P.C., 2003. The struggle to govern the commons. Science 302 (5652), 1907–1912. Doll, R., Hill, A.B., 1956. Lung cancer and other causes of death in relation to smoking. British Medical Journal 2 (5001), 1071. Ellis, J., 2010. The Effects of Fossil-Fuel Subsidy Reform: A Review of Modeling and Empirical Studies. International Institute for Sustainable Development, Manitoba. Engel, S., Pagiola, S., Wunder, S., 2008. Designing payments for environmental services in theory and practice: an overview of the issues. Ecological Economics 65 (4), 663–674. Ferraro, P.J., 2008. Asymmetric information and contract design for payments for environmental services. Ecological Economics 65 (4), 810–821. Ferrer-i-Carbonell, A., Gowdy, J.M., 2007. Environmental degradation and happiness. Ecological Economics 60 (3), 509–516. Fiala, N., 2008. Meeting the demand: an estimation of potential future greenhouse gas emissions from meat production. Ecological Economics 67 (3), 412–419. Fischer, C., 2001. Rebating Environmental Policy Revenues: Output-Based Allocations and Tradable Performance Standards. Resources for the Future, Washington, DC, pp. 1–22. Fischer, C., Fox, A.K., 2007. Output-based allocation of emissions permits for mitigating tax and trade interactions. Land Economics 83 (4), 575–599. Fisher, A.D., 1981. Exhaustible resources: the theory of optimal depletion. In: Resource and Environmental Economics. Cambridge University Press, New York. Fredriksson, P.G., Sterner, T., 2005. The political economy of refunded emissions payment programs. Economics Letters 87 (1), 113–119. Fullerton, D., 1997. Environmental levies and distortionary taxation: comment. The American Economic Review 87 (1), 245–251. Fullerton, D., 2008. Distributional Effects of Environmental and Energy Policy: An Introduction. National Bureau of Economic Research (No. w14241). Gersbach, H., Requate, T., 2004. Emission taxes and optimal refunding schemes. Journal of Public Economics 88 (3), 713–725. Gianessi, L.P., Peskin, H.M., Wolff, E., 1979. The distributional effects of uniform air pollution policy in the United States. The Quarterly Journal of Economics 93, 281–301.

ARTICLE IN PRESS References

Gilliland, P.M., Laffoley, D., 2008. Key elements and steps in the process of developing ecosystem-based marine spatial planning. Marine Policy 32 (5), 787–796. Gollier, C., Jullien, B., Treich, N., 2000. Scientific progress and irreversibility: an economic interpretation of the ‘Precautionary Principle’. Journal of Public Economics 75 (2), 229–253. Goulder, L., 1995. Environmental taxation and the double dividend: a reader’s guide. International Tax and Public Finance 2 (2), 157–183. Goulder, L.H., Parry, I.W.H., Burtraw, D., 1997. Revenue-raising vs. other approaches to environmental protection: the critical significance of pre-existing tax distortion. The Rand Journal of Economics 28 (4), 708–731. Goulder, L.H., Parry, I.W.H., Williams, R.C., Burtraw, D., 1999. The cost-effectiveness of alternative instruments for environmental protection in a second-best setting. Journal of Public Economics 72 (3), 329–360. Green, J., Sterner, T., Wagner, G., 2014. A balance of ‘bottom-up’ and ‘top-down’ in linking climate policies. Nature Climate Change 4, 1064–1067. https://doi.org/10.1038/NCLIMATE2429. Groom, B., Hill, D., Karousakis, K., Salzman, J., Sterner, T., Whitten, S., 2014. Biodiversity Offsets: Pros, Cons and Practical Issues. UNEP Policy Series. Gursoy, O., 2016. Do capital requirements in Basel III restrict the financing of green economy? A case study of a Turkish Bank. In: Social and Economic Perspectives on Sustainability. IJOPEC. Hagem, C., Holtsmark, B., Sterner, T., 2012. Mechanism Design for Refunding Emissions Payments. Statistics Norway, Oslo. Hahn, R.W., Stavins, R.N., 1991. Incentive-based environmental regulation: a new era from an old idea. Ecology L.Q. 18, 1. Hammar, H., Sterner, T., Åkerfeldt, S., 2013. Sweden’s CO2 tax and taxation reform experiences. In: Genevey, R., Pachauri, R., Tubiana, L. (Eds.), Reducing Inequalities: A Sustainable Development Challenge. TERI Press, New Delhi. Hanemann, M., 2014. Property rights and sustainable irrigation – a developed world perspective. Agricultural Water Management 145, 5–22. Hanley, N., Shogren, J.F., White, B., 1997. An economic analysis of non-renewable resources. In: Environmental Economics in Theory and Practice. Oxford University Press, Inc., New York. Heltberg, R., 2002. Property rights and natural resource management in developing countries. Journal of Economic Surveys 16 (2), 189–214. Henry, C., Tubiana, L., 2017. Earth at Risk. Columbia University Press. ISBN 9780231162524. Hepburn, C., 2010. Environmental policy, government, and the market. Oxford Review of Economic Policy 26 (2), 117–136. Hoel, M., 1991. Global environmental-problems – the effects of unilateral actions taken by one country. Journal of Environmental Economics and Management 20, 55–70. Hoel, M., 1998. Emission taxes versus other environmental policies. Scandinavian Journal of Economics 100 (1), 79–104. Hoel, M., Sterner, T., 2007. Discounting and relative prices. Climatic Change 84, 265–280. Hoffman, A.J., 2001. From Heresy to Dogma: An Institutional History of Corporate Environmentalism. Stanford University Press. Hotelling, H., 1931. The economics of exhaustible resources. Journal of Political Economy 39 (2), 137–175. IEA, 2016. Energy subsidies. International Energy Agency. https://www.iea.org/statistics/resources/ energysubsidies/. IMF, 2016. IMF Working Paper – How Large are Global Energy Subsidies? WP/15/105. International Monetary Fund. IPCC, 2014. Climate Change 2014: Synthesis Report. Contribution of Working Groups I, II and III to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. [Core Writing Team, Pachauri, R.K., Meyer, L.A. (Eds.)]. IPCC, Geneva, Switzerland. 151 pp. Jaeger, W.K., 2011. The welfare effects of environmental taxation. Environmental & Resource Economics 49 (1), 101–119.

49

ARTICLE IN PRESS 50

Selection and design of environmental policy instruments

Jaeger, W.K., 2012. The double dividend debate. In: Handbook of Research in Environmental Taxation. Edward Elgar Publishing (Chapter 12). Jaffe, A.B., Newell, R.G., Stavins, R.N., 2005. A tale of two market failures: technology and environmental policy. Ecological Economics 54 (2–3), 164–174. Jensen, S., Mohlin, K., Pittel, K., Sterner, T., 2015. An introduction to the green paradox: the unintended consequences of climate policies. Review of Environmental Economics and Policy 9 (2), 246–265. Johansson-Stenman, O., Brekke, K.A., 2008. The behavioral economics of climate change. Oxford Review of Economic Policy 24 (2), 280–297. Johansson-Stenman, O., Konow, J., 2010. Fair air: distributional justice and environmental economics. Environmental & Resource Economics 46, 147–166. Jung, C., Krutilla, K., Boyd, R., 1996. Incentives for advanced pollution abatement technology at the industry level: an evaluation of policy alternatives. Journal of Environmental Economics and Management 30 (1), 95–111. Kallbekken, S., Kroll, S., Cherry, T.L., 2011. Do you not like Pigou, or do you not understand him? Tax aversion and revenue recycling in the lab. Journal of Environmental Economics and Management 62, 53–64. Kennedy, P.W., 1994. Equilibrium pollution taxes in open economies with imperfect competition. Journal of Environmental Economics and Management 27 (1), 49–63. Keohane, N.O., Revesz, R.L., Stavins, R.N., 1998. The choice of regulatory instruments in environmental policy. Harvard Environmental Law Review 22, 313. Khalilian, S., Froese, R., Proelss, A., Requate, T., 2010. Designed for failure: a critique of the Common Fisheries Policy of the European Union. Marine Policy 34 (6), 1178–1182. Kitsakis, D., Dimopoulou, E., 2014. 3D cadastres: legal approaches and necessary reforms. Survey Review 46 (338), 322–332. Klenert, D., Mattauch, L., Combet, E., Edenhofer, O., Hepburn, C., Rafaty, R., Stern, N., 2017. Making Carbon Pricing Work. MPRA Paper No. 80943, posted 26 August 2017. Kling, C., Rubin, J., 1997. Bankable permits for the control of environmental pollution. Journal of Public Economics 64 (1), 101–115. Krutilla, K., 1991. Environmental regulation in an open economy. Journal of Environmental Economics and Management 20 (2), 127–142. Krutilla, K., Krause, R., 2011. Transaction costs and environmental policy: an assessment framework and literature review. International Review of Environmental and Resource Economics 4 (3–4), 261–354. Laestadius, L.I., Neff, R.A., Barry, C.L., Frattaroli, S., 2014. “We don’t tell people what to do”: an examination of the factors influencing NGO decisions to campaign for reduced meat consumption in light of climate change. Global Environmental Change 29, 32–40. Laffont, J.J., Tirole, J.A., 1993. Theory of Incentives in Procurement and Regulation. MIT Press, Cambridge, MA. Leach, M., Mearns, R., Scoones, I., 1999. Environmental entitlements: dynamics and institutions in community-based natural resource management. World Development 27 (2), 225–247. Levin, S., Xepapadeas, T., Crépin, A.S., Norberg, J., De Zeeuw, A., Folke, C., Hughes, T., Arrow, K., Barrett, S., Daily, G., Ehrlich, P., 2013. Social-ecological systems as complex adaptive systems: modeling and policy implications. Environment and Development Economics 18 (2), 111–132. Li, S., Linn, J., Muehlegger, E., 2014. Gasoline taxes and consumer behavior. American Economic Journal: Economic Policy 6 (4), 302–342. Libecap, G.D., 1978. Economic variables and the development of the law: the case of western mineral rights. The Journal of Economic History 38 (2), 338–362. Lin, B., Li, X., 2011. The effect of carbon tax on per capita CO2 emissions. Energy Policy 39 (9), 5137–5146. Lockie, S., 2013. Market instruments, ecosystem services, and property rights: assumptions and conditions for sustained social and ecological benefits. Land Use Policy 31, 90–98. Lowder, T., 2012. Should renewable energy be afraid of Basel III banking standards? Renewable Energy World. https://www.renewableenergyworld.com/articles/2012/08/should-renewable-energy-be-afraidof-basel-iii-banking-standards.html.

ARTICLE IN PRESS References

Maler, K., Fisher, A., 2005. Environment, uncertainty and option values. In: Handbook of Environmental Economics, vol. 2, pp. 571–620. Markussen, P., Svendsen, G.T., 2005. Industry lobbying and the political economy of GHG trade in the European Union. Energy Policy 33 (2), 245–255. Marteau, T.M., Ogilvie, D., Roland, M., Suhrcke, M., Kelly, M.P., 2011. Judging nudging: can nudging improve population health. British Medical Journal (online) 342, 263. Martin, R., Muûls, M., De Preux, L., Wagner, U., 2014. Industry compensation under relocation risk: a firm-level analysis of the EU emissions trading scheme. The American Economic Review 104, 2482–2508. McConnell, A., 2015. What is policy failure? A primer to help navigate the maze. Public Policy and Administration 30 (3–4). McKenney, B.A., Kiesecker, J.M., 2010. Policy development for biodiversity offsets: a review of offset frameworks. Environmental Management 45 (1), 165–176. Millennium Ecosystem Assessment, 2005. Ecosystems and Human Well-being: Synthesis. Island Press, Washington, DC. Millock, K.C., Nauges, C., Sterner, T., 2004. Environmental Taxes: A Comparison of French and Swedish Experience from Taxes on Industrial Air Pollution. CESifo DICE Report 1: 30–34. Mills, E., 2005. Insurance in a climate of change. Science 12, 1040–1044. Milman, O., Leavenworth, S., 2016. China’s plan to cut meat consumption by 50% cheered by climate campaigners. The Guardian 20. Miteva, D.A., Pattanayak, S.K., Ferraro, P.J., 2012. Evaluation of biodiversity policy instruments: what works and what doesn’t? Oxford Review of Economic Policy 28 (1), 69–92. Montero, J-P., 2002. Permits, standards and technology innovation. Journal of Environmental Economics and Management 44 (1), 23–44. Myers, N., 1998. Lifting the veil on perverse subsidies. Nature 392 (6674), 327. Naess, A., 1989. Ecology, Community and Lifestyle. Cambridge University Press. Nature Editorial, 2017. US Republican idea for tax on carbon makes climate sense. Nature 542, 271–272. Naturvårdsverket (Swedish EPA) 2004. Utvärdering av styrmedel i klimatpolitiken, Delrapport 2 i Energimyndighetens och Naturvårdsverkets underlag till Kontrollstation 2004. Nordhaus, W., 1974. Resources as a constraint on growth. The American Economic Review 64 (2), 22–26. Nordhaus, W., 2015. Climate clubs: overcoming free-riding in international climate policy. The American Economic Review 105, 1339–1370. Nyborg, K., 2016. Balladen om den usynlige hånd (The Ballad of the Invisible Hand). Short stories. Aschehoug, Oslo. Nyborg, K., 2017. Humans in the Perfectly Competitive Market: a Fictional Field Study. Nyborg, K., Anderies, J.M., Dannenberg, A., Lindahl, T., Schill, C., Schlueter, J., Adger, W.N., Arrow, K.J., Barrett, S., Carpenter, S., Chapin III, F.S., Crepin, A-S., Daily, G., Ehrlich, P., Folke, C., Jager, W., Kautsky, N., Levin, S.A., Madsen, O.J., Polasky, S., Scheffer, M., Walker, B., Wilen, J., Xepapadeas, A., de Zeeuw, A., 2016. Social norms as solutions. Science 354 (6308), 42–43. Olson, M., 1965. The Logic of Collective Action: Public Goods and the Theory of Groups. Harvard University Press. Oreskes, N., Conway, E.M., 2010. Defeating the merchants of doubt. Nature 465 (7299), 686. Ostrom, E., 1990. Governing the Commons: The Evolution of Institutions for Collective Action. Cambridge University Press, New York. Ostrom, E., 1998. A behavioral approach to the rational choice theory of collective action: Presidential address, American Political Science Association, 1997. American Political Science Review 92 (1), 1–22. Ostrom, E., 2000. Collective action and the evolution of social norms. Journal of Economic Policy 14, 137–158. Ostrom, E., 2005. Understanding Institutional Diversity. Princeton University Press, Princeton, NJ. Ostrom, E., 2009. Beyond Markets and States: Polycentric Governance of Complex Economic Systems. Nobel Prize lecture.

51

ARTICLE IN PRESS 52

Selection and design of environmental policy instruments

Ostrom, E., Burger, J., Field, C.B., Norgaard, R.B., Policansky, D., 1999. Revisiting the commons: local lessons, global challenges. Science 284 (5412), 278–282. Parry, I.W.H., Small, K.A., 2005. Does Britain or the United States have the right gasoline tax? The American Economic Review 95 (4), 1276–1289. Pearce, D., 1991. The role of carbon taxes in adjusting to global warming. The Economic Journal 101 (407), 938–948. Peltzman, S., 1976. Toward a more general theory of regulation. The Journal of Law & Economics 19 (2), 211–240. Pezzey, J.C.V., Park, A., 1998. Reflections on the double dividend debate. Environmental & Resource Economics 11 (3–4), 539–555. Pittel, K., Rübbelke, D.T., 2008. Climate policy and ancillary benefits: a survey and integration into the modelling of international negotiations on climate change. Ecological Economics 68 (1–2), 210–220. Pizer, W.A., 2002. Combining price and quantity controls to mitigate global climate change. Journal of Public Economics 85 (3), 409–434. Posner, R.A., 1974. Theories of economic regulation. The Bell Journal of Economics and Management Science 5, 335. Pulver, S., 2007. Making sense of corporate environmentalism: an environmental contestation approach to analyzing the causes and consequences of the climate change policy split in the oil industry. Organization & Environment 20 (1), 44–83. Rivers, N., Schaufele, B., 2015. Salience of carbon taxes in the gasoline market. Journal of Environmental Economics and Management 74, 23–36. Robinson, E.J., Albers, H.J., Busby, G.M., 2013. The impact of buffer zone size and management on illegal extraction, park protection, and enforcement. Ecological Economics 92, 96–103. Rockström, J., Steffen, W., Noone, K., Persson, A., Chapin III, F.S., Lambin, E.F., Lenton, T.M., Scheffer, M., Folke, C., Schellnhuber, H.J., Nykvist, B., de Wit, C.A., Hughes, T., van der Leeuw, S., Rodhe, H., Soerlin, S., Snyder, P.K., Costanza, R., Svedin, U., Falkkenmark, M., Karlberg, L., Corell, R.W., Fabry, V.J., Hansen, J., Walker, B., Liverman, D., Richardson, K., Crutzen, P., Foley, J.A., 2009. A safe operating space for humanity. Nature 461, 472–475. Rodrik, D., 2017. Populism and the economics of globalization. Draft PDF. https://drodrik.scholar.harvard. edu/publications/populism-and-economics-globalization. Rothstein, B., 2011. The Quality of Government: Corruption, Social Trust and Inequality in International Perspective. The University of Chicago Press, Chicago and London. Rubin, J.D., 1996. A model of intertemporal emission trading, banking and borrowing. Journal of Environmental Economics and Management 31 (3), 269–286. Sanchirico, J.N., Wilen, J.E., 2005. Optimal spatial management of renewable resources: matching policy scope to ecosystem scale. Journal of Environmental Economics and Management 50 (1), 23–46. Sanchirico, J., Wilen, J., 2007. Global Marine fisheries resources: status and prospects. International Journal of Global Environmental Issues 7 (2/3), 106–118. Schlager, E., Ostrom, E., 1992. Property-rights regimes and natural resources: a conceptual analysis. Land Economics 68, 249–262. Schmalensee, R., Stavins, R.N., 2013. The SO2 allowance trading system: the ironic history of a grand policy experiment. The Journal of Economic Perspectives 27 (1), 103–122. Simpson, R.D., 1995. Optimal pollution taxation in a Cournot duopoly. Environmental & Resource Economics 6 (4), 359–369. Sinn, H.W., 2015. The green paradox: a supply-side view of the climate problem. Review of Environmental Economics and Policy 9 (2), 239–245. Sjöstedt, M., Jagers, S.C., 2014. Democracy and the environment revisited: the case of African fisheries. Marine Policy 43, 143–148. Slunge, D., Sterner, T., 2001. Implementation of Policy Instruments for Chlorinated Solvents. European Environment 11 (5), 281–296. Smith, A., 1776. The Wealth of Nations: A Translation into Modern English. Industrial Systems Research. ISBN 978-0-906321-70-6.

ARTICLE IN PRESS References

Smith, V.K., Wolloh, C.V., 2012. Has Surface Water Quality Improved Since the Clean Water Act? (No. w18192). National Bureau of Economic Research. http://www.nber.org/papers/w18192. Somanathan, E., Sterner, T., Sugiyama, T., Chimanikire, D., Dubash, N.K., Essandoh-Yeddu, J.K., Fifita, S., Goulder, L., Jaffe, A., Labandeira, X., Managi, S., Mitchell, C., Montero, J.P., Teng, F., Zylicz, T., 2014. National and sub-national policies and institutions. In: Edenhofer, O., Pichs-Madruga, R., Sokona, Y., Farahani, E., Kadner, S., Seyboth, K., Adler, A., Baum, I., Brunner, S., Eickemeier, P., Kriemann, B., Savolainen, J., Schlömer, S., von Stechow, C., Zwickel, T., Minx, J.C. (Eds.), Climate Change 2014: Mitigation of Climate Change. Contribution of Working Group III to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA. Stavins, R.N., 1995. Transaction costs and tradeable permits. Journal of Environmental Economics and Management 29 (2), 133–148. Stavins, R.N., 1996. Correlated uncertainty and policy instrument choice. Journal of Environmental Economics and Management 30 (2), 218–232. Stavins, R.N., 2003. Experience with market-based environmental policy instruments. In: Handbook of Environmental Economics. Elsevier, pp. 355–435. Stavins, R.N., 2016. Market Mechanisms in the Paris Climate Agreement: International Linkage Under Article 6.2. The Paris Agreement and Beyond: International Climate Change Policy Post-2020. p. 53. Steffen, W., Richardson, K., Rockstroem, J., Cornell, S.E., Fetzer, I., Bennett, E.M., Biggs, R., Carpenter, S.R., de Vries, W., de Wit, C.A., Folke, C., Gerten, D., Heinke, J., Mace, G.M., Persson, L.M., Ramanathan, V., Reyers, B., Soerlin, V., 2015. Planetary boundaries: guiding human development on a changing planet. Science 347, 1259855. Sterner, T., 2007. Fuel taxes: an important instrument for climate policy. Energy Policy 35 (6), 3194–3202. Sterner, T., Coria, J., 2012. Policy Instruments for Environmental and Natural Resource Management. RFF Press, Routledge, Washington D.C. Sterner, T., Isaksson, L.H., 2006. Refunded emissions payments theory, distribution of costs, and Swedish experience of NOx abatement. Ecological Economics 57 (1), 93–106. Sterner, T., Köhlin, G., 2015. Pricing carbon: the challenges. In: Barrett, S., Carraro, C., De Melo, J. (Eds.), Towards a Workable and Effective Climate Regime. International Center for Climate Governance, p. 251. Sterner, T., Somanathan, E., 2006. Environmental policy instruments and institutions in developing countries. In: Toman, M., Lopez, R. (Eds.), Economic Development & Environmental Sustainability. Oxford University Press. Foreword by Joseph Stiglitz. Stigler, G.J., 1971. The theory of economic regulation. The Bell Journal of Economics and Management Science 2 (1), 3–21. Tietenberg, T., 2003. The tradable-permits approach to protecting the commons: lessons for climate change. Oxford Review of Economic Policy 19 (3), 400–419. Toman, M.A., Shogren, J.F., 2010. Climate change policy. In: Public Policies for Environmental Protection. Routledge, pp. 135–178. Ulph, A., 1992. The choice of environmental policy instruments and strategic international trade. In: Conflicts and Cooperation in Managing Environmental Resources. Springer, Berlin, Heidelberg, pp. 111–132. Ulph, A., 1998. Political institutions and the design of environmental policy in a federal system with asymmetric information. European Economic Review 42 (3–5), 583–592. Ulph, A., 2000. Harmonization and optimal environmental policy in a federal system with asymmetric information. Journal of Environmental Economics and Management 39 (2), 224–241. van der Horst, D., 2007. Assessing the efficiency gains of improved spatial targeting of policy interventions: the example of an agri-environmental scheme. Journal of Environmental Management 85 (4), 1076–1087. Verhoef, E., Nijkamp, P., Rietveld, P., 1997. Tradeable permits: their potential in the regulation of road transport externalities. Environment & Planning B, Planning & Design 24 (4), 527–548. Vogler, J., 2012. Global commons revisited. Global Policy 3 (1), 61–71.

53

ARTICLE IN PRESS 54

Selection and design of environmental policy instruments

Ward, H., Cao, X., 2012. Domestic and international influences on green taxation. Comparative Political Studies 45 (9), 1075–1103. Ward, J., Sammon, P., Dundas, G., Peszko, G., Kennedy, P.M., Wienges, S., Prytz, N., 2015. Carbon Leakage: Theory, Evidence, and Policy Design (English). Partnership for Market Readiness technical note; no. 11. World Bank Group, Washington, D.C. Watts, N., Adger, W.N., Agnolucci, P., Blackstock, J., Byass, P., Cai, W., Chaytor, S., Colbourn, T., Collins, M., Cooper, A., Cox, P.M., Depledge, J., Drummond, P., Ekins, P., Galaz, V., Grace, D., Graham, H., Grubb, M., Haines, A., Hamilton, I., Hunter, A., Jiang, X., Li, M., Kelman, I., Liang, L., Lott, M., Lowe, R., Luo, Y., Mace, G., Maslin, M., Nilsson, M., Oreszczyn, T., Pye, S., Quinn, T., Svensdotter, M., Venevsky, S., Warner, K., Xu, B., Yang, J., Yin, Y., Yu, C., Zhang, Q., Gong, P., Montgomery, H., Costello, A., 2015. Health and climate change: policy responses to protect public health. The Lancet 386 (10006), 1861–1914. https://doi.org/10.1016/s0140-6736(15)60854-6. Watts, N., Amann, M., Ayeb-Karlsson, S., Belesova, K., Bouley, T., Boykoff, M., Byass, P., Cai, W., Campbell-Lendrum, D., Chambers, J., Cox, P.M., 2018. The Lancet Countdown on health and climate change: from 25 years of inaction to a global transformation for public health. The Lancet 10120 (10–16). 540 pp. Wätzold, F., Schwerdtner, K., 2005. Why be wasteful when preserving a valuable resource? A review article on the cost-effectiveness of European biodiversity conservation policy. Biological Conservation 123 (3), 327–338. Weitzman, M.S., 1974. Prices vs quantities. The Review of Economic Studies 41 (4), 477–491. Weitzman, M.L., 2017. On a world climate assembly and the social cost of carbon. Economica 84 (336), 559–586. Welsch, H., 2006. Environment and happiness: valuation of air pollution using life satisfaction data. Ecological Economics 58 (4), 801–813. Welsch, H., 2009. Implications of happiness research for environmental economics. Ecological Economics 68 (11), 2735–2742. West, S.E., Williams III, R.C., 2007. Optimal taxation and cross-price effects on labor supply: estimates of the optimal gas tax. Journal of Public Economics 91, 593–617. Wigley, T.M.L., Richels, R., Edmonds, J.A., 1996. Economic and environmental choices in the stabilization of atmospheric CO2 concentrations. Nature 379 (6562), 240. Woerdman, E., Arcuri, A., Clò, S., 2008. Emissions trading and the polluter-pays principle: do polluters pay under grandfathering? Review of Law & Economics 4 (2), 565–590. Wordie, J.R., 1983. The chronology of English enclosure, 1500–1914. The Economic History Review 36 (4), 483–505. World Bank, Ecofys, 2017. Carbon Pricing Watch 2017. An Advance Brief from the “State and Trends of Carbon Pricing 2017” report, to be released late 2017. World Bank, Ecofys, Vivid Economics, 2016. State and Trends of Carbon Pricing. World Bank, Washington, DC. © World Bank. https://openknowledge.worldbank.org/handle/10986/25160. License: CC BY 3.0 IGO. Wu, J., Boggess, W.G., 1999. The optimal allocation of conservation funds. Journal of Environmental Economics and Management 38 (3), 302–321. Wunder, S., 2005. Payments for environmental services: some nuts and bolts. Wunder, S., Wertz-Kanounnikoff, S., 2009. Payments for ecosystem services: a new way of conserving biodiversity in forests. Journal of Sustainable Forestry 28 (3–5), 576–596. Wünscher, T., Engel, S., Wunder, S., 2008. Spatial targeting of payments for environmental services: a tool for boosting conservation benefits. Ecological Economics 65 (4), 822–833. Xepapadeas, A.P., 1992. Environmental policy design and dynamic nonpoint-source pollution. Journal of Environmental Economics and Management 23 (1), 22–39.