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Forthcoming in Oxford Handbook of Law and Economics (2014)

Draft of January 24, 2014

Environmental Law and Economics

Michael A. Livermore & Richard L. Revesz

1. INTRODUCTION ..................................................................................................................................................... 1

2. COST-BENEFIT ANALYSIS ................................................................................................................................... 2

2.1. NORMATIVE ISSUES AND ANALYSIS ................................................................................................................................... 3 2.1.1 General Framework ................................................................................................................................................. 3 2.1.2 Discounting and future generations .................................................................................................................. 4 2.1.3 Employment effects .................................................................................................................................................. 6 2.1.4 Behavioral economics and the energy paradox ............................................................................................ 7

2.2. POSITIVE ISSUES AND ANALYSIS ........................................................................................................................................ 8 2.2.1. A changing political context for cost-benefit analysis............................................................................... 8 2.2.2 Social cost of carbon ............................................................................................................................................. 12

3. INSTRUMENT CHOICE ....................................................................................................................................... 13

3.1 NORMATIVE ISSUES AND ANALYSIS .................................................................................................................................. 13 3.1.1 General Framework .............................................................................................................................................. 13 3.1.2 Market-mechanisms vs. command-and-control ........................................................................................ 14 3.1.3 Environmental transitions and grandfathering ........................................................................................ 15

3.2 POSITIVE ISSUE AND ANALYSIS ......................................................................................................................................... 16 3.2.1 Experience with market-based mechanisms ............................................................................................... 16 3.2.2 Labeling and disclosure ....................................................................................................................................... 17 3.2.3 Social norms ............................................................................................................................................................. 18 3.2.4 Reducing the effect of grandfathering under the Clean Air Act .......................................................... 19

4. JURISDICTIONAL ALLOCATION ..................................................................................................................... 19

3.1 NORMATIVE ISSUES AND ANALYSIS .................................................................................................................................. 19 3.2 POSITIVE ISSUES AND ANALYSIS ....................................................................................................................................... 20

3.2.1 Jurisdictional mismatch ...................................................................................................................................... 20 3.2.2 Lack of a global agreement on climate change ......................................................................................... 21 3.2.3 State innovation ..................................................................................................................................................... 21

5. CONCLUSION ........................................................................................................................................................ 23

1. Introduction

Law and economics provides a useful lens for many environmental policy

questions. Normative deliberation concerning the construction of environmental policy

can be informed by an economics perspective. Economic analysis can also be brought to

bear on empirical questions concerning the effects of environmental policies (Faure

2012) and the political economy factors that affect the selection of environmental policies

(Keohane, Revesz, Stavins 1996; Burtraw 2012). Indeed, the overlap between

environmental policy and economics is sufficiently extensive that environmental

economics constitutes a distinct discipline within the field of economics (Field and Field

2012).

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Revesz and Stavins (2007) presented an overview of environmental law and

economics, focusing on three themes that have been of particular importance in the field:

cost-benefit analysis, instrument choice, and the allocation of policymaking responsibility

in a federal system. That piece canvassed the robust literature that had developed on

those themes and examined the political and policy contexts that affected the application

of law and economics to environmental policymaking.

There have been important developments since the publication of that piece.

Climate change has become an even more dominant issue in environmental law, both in

the United States and internationally. The scale of the threatened harms from climate

disruption, the global and diffused nature of greenhouse gas emissions, the degree of

scientific uncertainty, the long time-horizons involved, and the extent of economic

transformation needed to substantially curtail emissions all make climate change a

uniquely vexing environmental problem (Intergovernmental Panel on Climate Change

[IPCC] 2013)).

The issue of climate change has precipitated important changes in the field of

environmental law and economics. Many normative and positive challenges are posed by

climate change and attempts to reduce greenhouse gas emissions, and scholars have

responded to those challenges with important conceptual and empirical advances. The

political context of environmental law and economics has also been affected by conflict

over climate policies. After providing a brief general overview of the economics of

environmental law, this chapter will focus on these recent developments, with an

emphasis on the experience of the United States. Consistent with Revesz and Stavins

(2007), this chapter focuses on pollution control and does not examine natural resource

management.

When setting environmental policy, decision makers must address two general

types of questions. The first concerns the ends of environmental policy, and examines the

socially desirable level of environmental quality. The second type of question concerns

the means of policy making and focuses on the types of regulatory instruments that will

be used and the allocation of responsibility between governmental actors. Section 2

addresses the first type of question concerning the goals of environmental policy.

Sections 3 and 4 address the means of environmental policy, focusing on instrument

choice and jurisdictional allocation, respectively.

2. Cost-benefit analysis

There is widespread agreement that environmental quality is a good that is worth

providing. Not only do people enjoy the benefits of environmental quality directly—for

example through improved health and better recreational opportunities—but natural

systems provide the foundation for many diverse forms of economic activity (such as

agriculture) and, indeed, create the basic conditions for human life to exist at all. But

although it is impossible to dismiss the value of a clean environment, there is much less

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agreement on the level of environmental quality that society ought to pursue. The

question of “how clean is clean enough?” is an essential preliminary to environmental

policy making, and one on which economics provides useful guidance.

2.1. Normative Issues and Analysis

Welfare economic provides a general framework for answering questions

concerning the ends of environmental policy. The general recommendation is that

environmental policy should be selected to maximize well-being, measured by the value

that individuals place on improved environmental quality minus the value of sacrifices

that must be made to achieve those improvements. Because most environmental policy

involves tradeoffs between positive and negative consequences, the value of those

consequences must be weighed against each other. The dominant technique for weighing

policy consequences and estimating the net effects of policy options is cost-benefit

analysis.

2.1.1 General Framework

Pollution presents a classic market failure that can be addressed through

appropriate government intervention. Pollution can be described as an externality (Pigou

1920) in which the “activity of one agent . . . affects the well-being of another agent and

occurs outside the market mechanism” (National Research Council 2010). The solution

proposed by Pigou is a tax or fee to internalize the external costs imposed by pollution.

Such a move is not necessary in world of well-defined property rights and low transaction

costs because private bargaining will account for all relevant effects (Coase 1960;

Krutilla and Krause 2011). Nevertheless, the existence of transaction costs and

incomplete property rights imply that private bargaining alone will not always results in a

social welfare maximizing outcome.

Once it is determined that government intervention can, in principle, improve

well-being, the question of the socially desirable level of pollution control must be

addressed. Within the field of welfare economics, the traditional answer is the test

developed by Kaldor (1939) and Hicks (1939) of potential Pareto improvements. A

Pareto improvement is one that makes at least one person better off and no one worse off

(Pareto 1896). A potentially Pareto improvement is one in which the beneficiaries of the

policy could fully compensate those who are burdened by the policy.

Kaldor-Hicks efficiency is the basis for formal cost-benefit analysis (Zerbe and

Bellas 2006). A measure is Kaldor-Hicks efficient if it maximizes the difference between

the value of its benefits (as measured by the beneficiaries) and its costs (as measured by

bearers of those costs). Because total benefits typically increase at a decreasing rate while

cost increase at an increasing rate, net benefits are maximized by policies that equate

marginal benefits with marginal costs.

There is a robust debate over the normative attractiveness of cost-benefit analysis

as a means of identifying socially desirable environmental policy. Issue that are relevant

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to this debate include the normative status of preference satisfaction (Adler and Posner

2006), the importance of distribution of costs and benefits for well-being (Adler 2012;

Cai et al. 2010; Farrow 2011), as well as broader critiques concerning

incommensurability (Anderson 1995; Ackerman & Heinzerling 2004) or

commodification (Radin 2001). While many of the criticisms leveled against cost-benefit

analysis apply generally to normative economics, the use of cost-benefit analysis to

evaluate environmental policies has been particularly controversial (Kysar 2010).

Additional conceptual frameworks for evaluating environmental policy include

environmental rights (Kelman 1981a), environmental justice (Rhodes 2003), non-

anthropocentric approaches (Ariansen 1998), green growth (Livermore 2013) and

sustainable development (Dasgupta and Heal 1974; Stiglitz 1974; Solow 1974;

Organisation for Economic Cooperation and Development [OECD] 2004). Many of

these concepts can be understood as emphasizing distributional, rather than efficiency,

considerations (Stavins et al 2003). For example, one definition of sustainable

development that is reasonably well-accepted in the economics community is

monotonically increasing consumption (Solow 1986; Hartwick 1977; Daly 1990), which

concerns the inter-generational allocation of resources, rather than the static

maximization of welfare.

Environmental policymaking poses several methodological difficulties for cost-

benefit analysis. Several of these are described in detail in Revesz and Stavins (2007).

The most important environment-specific challenge is to provide valid monetary

estimates of the value of hard-to-price goods, including non-market goods (like clean air)

(Scarpa and Alberini, eds. 2005; Banzhaf 2010; Phaneuf et al. 2009) and mortality risk

reduction (Cropper et al. 2011). Continent valuation (e.g., stated preference studies)

remains the most widespread tool to estimate non-market values (Viscusi 2008). While

this approach has been endorsed by leading economists, it remains controversial

(Hausman 2012). In recent years, the concept of ecosystem services—the goods and

services provided by natural systems—has been used to help structure research on the

benefits of environmental protection (Goulder and Kennedy 2011; Brown et al. 2007). An

important challenge for the ecosystems services approach is to identify estimation

methods that disaggregate intermediary goods, such as water quality, from final services,

such as recreation or human health (Keeler et al. 2012).

2.1.2 Discounting and future generations

Carbon dioxide, the most important greenhouse gas pollutant, has a long

atmospheric lifetime, on the order of one hundred years (IPCC 2013). In addition, global

warming processes are not instantaneous, but involve the interaction of large, complex

physical systems that may involve irreversible feedback loops (IPCC 2013). The

consequences of greenhouse gas emissions today, then, will be felt for many years into

the future. As in many other environmental contexts, limits on greenhouse gas emissions

impose immediate costs to produce benefits that will accrue only in the long term.

Traditionally, future benefits have been discounted at a constant rate. The choice of a

discounting procedure and rate is particularly important in the case of climate change

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because of the very long time horizons involved (Revesz and Shahabian 2011). The

standard approach to analyzing this problem is contained in the following formula

(Arrow et al. 1996):

d = ρ + θg

where d is the discount rate, ρ is the rate of pure time preference, θ is the absolute value

of the elasticity of marginal utility of consumption and g is the growth rate of per capita

consumption.

There are two interpretations for the first term, ρ. Under a prescriptive approach,

ρ would be derived from ethical principles (Arrow et al. 1996). There is no strong

consensus concerning whether a pure time preference is normatively justified. Many

prominent economists agree with the proposition, first articulated by Ramsey (1928), that

ρ should be zero (Broome 1992; Dasgupta 2008; Cline 2006; Harrod 1966; Heal 2009;

Koopmans 1967; Philibert 1999; Pigou 1932; Solow 1974). Others disagree (Arrow

1999; Beckerman & Hepburn 2007).

Revesz (1999) and Revesz and Shahabian (2011) argue that there is an ethical

distinction between inter-generational and intra-generational discounting that is relevant

for determining the appropriate pure rate of time preference. The latter reflects

individuals’ preferences to spread consumption across their lifetime. The former reflects

a social decision concerning the allocation of resources between individuals. While there

are reasons to endorse policies that reflect individuals’ desire to consume sooner rather

than later, there is no straightforward justification for a bare social preference for current

over future generations.

The descriptive approach avoids the preceding normative questions and simply

substitutes observed market interest rates for the discount rate in cost-benefit analysis.

The benefit of this approach is that it does not attempt to address difficult questions of

inter-generational equity. The problem is that it fails to provide a normative reason for

why market interest rates are an appropriate source for the social discount rate (Revesz

and Shahabian 2011). Sophisticated defenders of inter-generational discounting recognize

the possibility for “net welfare losses and distributional inequity” but argue that the issue

of moral obligations to future generations should be treated as a separate inquiry from

efficiency (Sunstein and Rowell 2007).

Once society decides how to trade off utility between generations, discounting can

be used to account for opportunity costs in determining the degree to which pollution

control should be included in the basket of future-oriented investments (Schelling 1995,

Samida and Weisbach 2007, Weisbach and Sunstein 2009).

Accounting for the marginal utility of consumption through growth discounting

also generates difficulties. As a preliminary matter, long-term growth is difficult to

predict (Moyer et al. 2013). The countries most likely to benefit in the future from current

climate change mitigation—developing countries in the tropics—are far poorer than the

industrial powers that are most likely to pay in the present for such mitigation (Ruhl

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2012). And even in the relatively distant future, these developing countries may be

poorer than developed countries are today. Scholars have also argued that greenhouse

gas reductions in developed countries can be viewed as foreign aid and noted that intra-

generational wealth transfers are a more efficient mechanism to reduce global inequality

than climate mitigation (Schelling 1995; Sunstein & Posner 2007).

There appears to be an emerging consensus that discount rates should decline with

the length of the time horizon, with higher discount rates applied to the near-term future

and lower discount rates applied to the long-term future (Gollier and Weitzman 2009). A

declining discount rate approach has been adopted by France and the United Kingdom

(Cropper 2012). The first argument in favor of such an approach is that it is more

consistent with observed behavior. Stated preference studies typically indicate that

individuals apply a very low discount rate to benefits to future generations (Cropper et al.

1992; Johannesson and Johansson 1996). Even in individual market behavior, hyperbolic

discounting is commonly observed (Laibson 1997), although it raises rationality concerns

(Skog 2005).

The second and more powerful argument in favor of declining discount rates

stems from uncertainty about the appropriate discount rates (Arrow et al. 2013; Gollier

and Weitzman 2009). When these rates are uncertain, the utility maximizing approach

requires that decisions be made based on expected discounted costs and benefits, not

discounted expected costs and benefits (Weitzman 1998, 2001). The consequence is that

the correct discount rate is lower than the mean of the probability distribution of possible

rates. Over sufficiently long time horizons, the lowest rate in the distribution dominates

(Weitzman 2010). Newell and Pizer (2003) show that rational treatment of interest rate

uncertainty, based on historic variance in U.S. Treasury bill rates, leads to quickly

declining discount rates.

2.1.3 Employment effects

A growing area of research interest, spurred in part by changes in the political

context of cost-benefit analysis discussed below, concerns the effects of environmental

regulation on employment. A threshold question is whether, and how, effects on labor

markets ought to be accounted for in cost-benefit analyses of environmental regulation.

The standard approach for government agencies in the past had been to assume a

well-functioning, full-employment labor market (EPA 2000). Under these conditions,

workers hired to comply with an environmental regulation are reallocated from positions

with similar wages, and workers that are laid off due to an environmental regulation find

new positions at similar compensation levels.

When the assumption of full employment is relaxed, the estimation of the labor

effects of environmental regulation is extremely difficult (Carrigan, Coglianese, and

Finkel 2014). Morgenstern, Pizer, and Shih (2002) note three distinct potential effects of

regulation on employment. A demand effect occurs if regulation increases production

costs, thereby increasing prices and lowering the quantity of production, resulting in

lower demand for labor. A cost-effect occurs if plants must add more capital and labor

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per unit of output, potentially increasing demand for labor. A factor-shift effect occurs if

a regulation induces a shift in spending from capital to labor or vice versa. The authors

note that the net effects of these three effects are ambiguous, and in a study of four highly

polluting and regulated industries, find mild positive employment effects from regulation.

Sectoral studies have found negative employment effects from environmental regulation,

but may measure employment shifts rather than net losses (Greenstone 2002; Kahn

1997). Walker (2011) examines wage effects from the Clean Air Act, finding that wages

within regulated industries declined as a consequence of environmental controls. OMB

(2011) provides a useful overview of research on employment effects from

environmental regulation.

Masur and Posner (2012) argue that, when realistic economic conditions are taken

into account, there can be substantial costs associated with layoffs. Not only do laid-off

workers incur transition costs, including job search costs and time away from work, but

there are psychological and physical hardships associated with joblessness that are not

fully captured by lost wages or search costs. Adler (2014) discusses how a welfare-

oriented approach would account for these types of harms.

Labor effects from environmental regulation are not always negative. If firms that

must hire new workers to comply with an environmental regulation draw from a pool of

unemployed labor, then the social costs are below the wages that are paid out to those

workers, because the opportunity costs of these workers are very low (Bartik 2013).

There may also be physical and psychological benefits associated with hiring

unemployed workers that imply a negative social cost (CBO 2012).

2.1.4 Behavioral economics and the energy paradox

Over the last two decades, scholars have begun applying the insights of

behavioral economics to the law (Jolls this volume [Ed note: not sure the style for these

types of internal citations]; Jolls et al. 1998; Krieger 1995; Korobkin and Ulen 2000;

McAdams 1997; Sunstein 1996). The behavioral revolution has touched environmental

economics as well (Shogren & Taylor 2008). Arguably the most important consequence

of behavioral research for cost-benefit analysis of environmental policy concerns the

issue of energy efficiency standards. Measures to promote energy efficiency have taken

on increasing importance as a low cost means of reducing greenhouse gas emissions.

Energy efficiency standards have two potential categories of benefits. By

reducing energy consumption, they can reduce externalities such as air pollution from

burning fossil fuels or thermal pollution associated with nuclear power generation. A

second category of benefits is consumer savings. This second category of benefits,

however, is complicated by the fact that rational consumers should be willing to invest in

all net present value positive energy efficiency improvements (Jaffe et al. 2004).

Therefore, under the standard model of rational economic behavior, it should be

impossible for government regulation to generate consumer benefits through energy

efficiency mandates in most circumstances.

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Notwithstanding this prediction, there is a large literature demonstrating

widespread underinvestment in energy efficiency (Gillingham et al. 2009). This

phenomenon is referred to as the “energy paradox” in the environmental economics

literature (Jaffe and Stavins 1994). There are several behavioral hypotheses that have

been forwarded to explain the energy paradox: consumers may myopically apply above

market discount rates to future energy savings (Alcott and Wozny 2012); consumers may

account for energy efficiency only after having settled on other product characteristics

(Geistfeld et al. 1977); consumers may demonstrate loss aversion (Greene 2008);

consumers may incorrectly associate energy efficiency with poor performance (Rogers

and Shrum 2012); or consumers may use other decision heuristics that fail to account for

efficiency (Helfand and Wolverton 2011). There is no consensus on the dominant source

of the energy paradox.

Despite the lack of a strong theoretical understanding of the energy paradox, its

persistence in the marketplace indicates that consumer savings are a valid benefit

associated with energy efficiency standards. While market-based tools, such as an energy

tax, are likely to be lower-cost tools for achieving emissions reductions (Karplus and

Paltsev 2012), energy efficiency standards may be justified even in situations where they

have no environmental benefits, for example when implemented together with a

comprehensive cap on emissions. In those cases, efficiency standards could be justified

purely on the basis of consumer savings (Bubb and Pildes 2014, forthcoming).

2.2. Positive Issues and Analysis

Cost-benefit analysis has become the dominant paradigm for environmental

policy analysis in the United States, and its use is now becoming more common around

the globe. But that does not mean that the technique is uncontroversial, and debates about

its use continue.

2.2.1. A changing political context for cost-benefit analysis

In U.S. environmental law, two general alternatives to cost-benefit analysis for

setting environmental standards are absolute standards (Sinden 2005) and feasibility

standards (Driesen 2005). Absolute standards seek to set environmental quality at levels

that present zero (or negligible) risk of harm. The National Ambient Air Quality

Standards (NAAQS) under the Clean Air Act, which must be set at a level “requisite to

protect the public health” with an “adequate margin of safety”1 are the most important

example of an absolute health-based standard in U.S. environmental law. Feasibility

standards are set to maximize stringency, subject to the constraints of “physically

impossible environmental improvements” or standards “so costly that they cause

widespread plant shutdowns” (Dreisen 2011). Both absolute standards and feasibility

standards are subject to serious objections. Absolute standards raise conceptual and

practical problems if there is no known threshold that presents zero risk (McGarity 1979;

1 42 U.S.C. § 7409(b)(1).

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Coglianese and Marchant 2004; Sunstein 1999). Furthermore, and paradoxically, the

majority of recent NAAQS (adopted under an absolute health-based approach) are less

stringent than would be economically efficient, which undermines the normative

justification for such an approach (Livermore and Revesz 2014, forthcoming). Feasibility

standards have been criticized as “creating significant problems of over- and

underregulation” because they are completely insensitive to some costs while weighing

other costs as infinitely high (Masur and Posner 2010).

Perhaps in part because of the limitations of these alternatives, cost-benefit

analysis has come to be the dominant framework for U.S. environmental policy (Sunstein

2002). While cost-benefit analysis is required in only a few environmental statutes ,

notably the Toxic Substances Control Act,2 the Federal Insecticide, Fungicide, and

Rodenticide Act,3 and the Safe Drinking Water Act,

4 for more than thirty years, U.S.

Presidents have required that significant proposed regulations be subject to cost-benefit

analysis (Zerbe, this volume). This presidential requirement applies to all agency

decisions unless precluded by statute.5 The most important case limiting the use of cost-

benefit analysis within the environmental arena is American Trucking Associations, Inc.

v. Whitman6 in which the Supreme Court held that the statutory silence in section 109 of

the Clean Air Act should be interpreted to bar the consideration of costs (and therefore

the use of cost-benefit analysis) for setting the NAAQS. The effect of American Trucking

on other statutory provisions was called into question in the 2009 decision Entergy Corp.

v. Riverkeeper, Inc.7 In that case, the Supreme Court declined to extend American

Trucking to cover a provision of the Clean Water Act that was silent on the use of cost-

benefit analysis. After Riverkeeper, it appears that, except in perhaps a handful of

instances, Executive Order 13,563 will apply and require agencies to justify their

significant environmental regulations by reference to cost-benefit analysis.

Cost-benefit analysis and the closely related practice of regulatory impact analysis

are now common practices outside the United States as well. The European Union has

undertaken several important steps to promote assessment of costs and benefits of

regulatory policy (Wiener 2006). Cost-benefit analyses carried out by governments,

independent academics, or non-governmental organizations is now commonly used in

many developing and emerging countries to evaluate environmental policy as well

(Livermore and Revesz 2013).

The global growth of cost-benefit analysis, and its universal support by Presidents

of both major U.S. political parties over the past three decades, masks a complex political

struggle over the degree to which concerns over economic efficiency ought to inform

how environmental policy is set. Because there are important policy consequences

2 15 U.S.C. §§ 2601–2629 (2006). 3 7 U.S.C. §§ 136–136y (2006). 4 42 U.S.C. § 300f et seq. (2006). 5 Executive Order 13563, Improving Regulation and Regulatory Review, 76 Fed. Reg. 3821 (Jan. 18, 2011). 6 531 U.S. 457 (2001). 7 556 U.S. 208 (2009).

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associated with cost-benefit analysis, its use is not only debated by scholars; organized

interest groups participate in these debates as well (Revesz and Livermore 2008).

In the United States, when the Reagan executive order placed cost-benefit

analysis at the center of the regulatory state, protection-oriented groups (such as

environmentalists) strongly opposed the move, while interest groups that favored more

lax regulation (such as industry trade associations) promoted expanded use of the

technique. This interest group dynamic remained remarkably consistent over time,

surviving multiple changes in party control of the White House (Revesz and Livermore

2008).

Recently, however, the political alignment around cost-benefit analysis has

become less stable. In light of congressional failure to enact climate legislation in 2009,

the Obama administration has pursued an aggressive agenda of clean air protections,

which either directly limit greenhouse gas emissions or have substantial climate co-

benefits. At the same time, the administration has continued the long-standing

presidential support for cost-benefit analysis and relied heavily on economic arguments

to support more stringent environmental standards.

Three environmental rules adopted by the EPA under the Obama administration

were supported by particularly persuasive cost-benefit analyses: the updated Corporate

Average Fuel Economy (CAFE) Standards released in 2012; the 2013 Mercury and Air

Toxics Standards (MATS); and EPA’s Cross State Air Pollution Rule. The CAFE

Standards will raise average fuel economy to 54.5 mpg by 2025, with estimated net

benefits of approximately $700 billion by 2025 (National Highway Traffic Safety

Administration 2012). The MATS extended air quality standards for mercury and other

toxic pollutants to power plants, with estimated net benefits of at least $27 billion in 2016

(Environmental Protection Agency [EPA] 2011a). The CSAPR is the EPA’s latest

attempt to address air quality problems presented by interstate externalities. The agency

estimates that the rule will produce at least $120 billion annually in net benefits by 2014

(EPA 2011b). (The CSAPR was struck down by the D.C. Circuit in 20128 and is

currently pending before the Supreme Court9). [Ed note: update if decision by print

date]. Statutorily mandated retrospective analysis prepared by the EPA of air quality

rules adopted pursuant to the Clean Air Act Amendments of 1990 estimated that by 2020,

the rules will have at least $1 trillion in net benefits, and, under more favorable

assumptions, up to $35 trillion (EPA 2011c).

Perhaps unsurprisingly, willingness to endorse cost-benefit analysis has shifted in

light of these developments. Protection-oriented groups have shown greater openness to

cost-benefit analysis (Livermore and Revesz 2011). At the same time, regulatory

skeptics have distanced themselves from the technique (Volokh 2011) as did the 2012

Republican presidential nominee Mitt Romney, who argued that cost-benefit analysis

“tend[s] to be vulnerable to manipulation and also disconnected from the central issue

8 EME Homer City Generation, L.P. v. EPA, 696 F.3d 7 (D.C. Cir. 2012). 9 EPA v. EME Homer City Generation, L.P., 133 S. Ct. 2857 (2013) (granting certiorari).

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confronting our country today, namely, generating economic growth and creating jobs”

(Romney 2011).

An important component of this political realignment is an expanded emphasis on

the employment effects of regulation by the regulated community and some politicians.

The phrase “job-killing regulation,” which appeared only four times in major U.S.

newspapers in 2007, appeared nearly 700 times in the same outlets four years later

(Livermore and Schwartz 2014). During the 112th congressional session, employment

effects were used to justify several bills to reduce EPA’s regulatory authority.10

The

House Republican Plan for America’s Job Creators included several provisions affecting

the rulemaking process (House Republican Conference 2011) and the Regulatory Freeze

for Jobs Act, which would block all new significant regulations until unemployment

dropped below 6 percent, passed the House in July 2012.11

Estimates of the employment effects of regulation have become common in

political discourse over environmental policy (Livermore and Schwartz 2014). These

estimates, which are typically based on input-output or computable general equilibrium

models, are highly sensitive to analytic assumptions and modeling choices, resulting in

widely disparate results. In one telling example, a report by the America Coalition for

Clean Coal Electricity estimated that two EPA rules would trigger 1.4 million job losses,

while a Political Economy Research Institute (PERI) study predicted the same two rules

would spur a 1.4 million job gain (Livermore and Schwartz 2014).

10 For example, the following bills were proposed during 2011 by the 112th Congress: H.R. 1 (a continuing appropriations resolution for FY2011, which passed the House in February, containing more than twenty riders restricting or prohibiting the use of funds to implement various regulatory activities under EPA’s jurisdiction); H.R. 199, Protect America’s Energy and Manufacturing Jobs Act of 2011 (proposing a two-year suspension of climate rules); H.R. 457, H.R. 517, H.R. 2018, & S. 272 (to modify EPA’s authority under the Clean Water Act); H.R. 750 & S. 228, Defending America’s Affordable Energy and Jobs Act (preempting any regulation to mitigate climate change); H.R. 872, Reducing Regulatory Burdens Act of 2011 (amending the Clean Water Act and FIFRA to alter EPA regulation of pesticide discharge into water); H.R. 910, Energy Tax Prevention Act (to prevent greenhouse gas regulations under the Clean Air Act); H.R. 960 & S. 468, Mining Jobs Protection Act (amending EPA’s consultation procedure under the Clean Water Act); H.R. 1391, Recycling Coal Combustion Residuals Accessibility Act of 2011, & H.R. 1405 (prohibiting coal ash from being regulated under Subtitle C of RCRA ); H.R. 2021, Jobs and Energy Permitt ing Act of 2011 (amending the Clean Air Act to change permitting of off shore sources); H.R. 2250 & S. 1392, EPA Regulatory Relief Act of 2011 (to delay the Boiler MACT rules); H.R. 2584 (an appropriations bill with various riders); H.R. 2681, Cement Sector Regulatory Relief Act (to delay the Cement MACT rules); H.R. 3400 & S. 1720, Jobs Through Growth Act (incorporating several of the above restrictions on EPA authority); and S.J. Res. 27 (a resolution to disapprove EPA’s cross-state air pollution rule). 11 H.R. 4078, Regulatory Freeze for Jobs Act (July 25, 2012). Available: http://www.gop.gov/bill/112/2/hr4078.

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Agencies have also begun to include employment effects alongside cost-benefit

analyses. The EPA in particular has included a statement on employment effects in

several high profile rulemakings (e.g., EPA 2012a; EPA 2012b). Employment effect

estimates generated by agencies are often quite modest, providing a useful counterweight

to extravagant jobs claims made by advocacy organizations.

2.2.2 Social cost of carbon

The application of cost-benefit analysis to regulations with greenhouse gas

emissions has become a particular focal point for political conflict over the methodology.

In 2009, the Obama administration convened an interagency task force that was charged

with developing a “social cost of carbon” to be used across the government to assign a

monetary value to greenhouse gas emission reductions in cost-benefit analyses of

rulemakings (Interagency Working Group on Social Cost of Carbon [IWG] 2010). The

original central estimate provided by the 2010 report for Year 2015 was $24 (2007

dollars), using a 3% discount rate (IWG 2010), and an updated estimate was released in

2013 of $38 (IWG 2013). Since being developed, the social cost of carbon has been used

in a number of important rulemakings, including EPA’s fuel-efficiency standards and

several Department of Energy appliance efficiency rules (e.g. EPA 2010a; Department of

Energy 2013).

The taskforce based its estimate on three Integrated Assessment Models (IAMs)

that link the physical and economic effects of climate change: the Dynamic Integrated

Model of Climate and the Economy (DICE) (Nordhaus and Sztorc 2013), the Climate

Framework for Uncertainty, Negotiation, and Distribution (FUND) model (Anthoff and

Tol 2013), and the Policy Analysis of the Greenhouse Effect (PAGE) model (Hope

2013). These models translate predictions concerning greenhouse gas emission,

temperature change, and physical impacts associated with climate change (such as sea

level rise or effects on agricultural productivity) into monetary terms.

Because many of the systems that are represented by these IAMs are not well

understood, there is considerable uncertainty concerning the accuracy of their estimates

(Pindyck 2013). Some scholars have argued that these uncertainties make climate change

a special case where cost-benefit analysis cannot be usefully applied (Masur and Posner

2011; Rose-Ackerman 2011). These arguments have been repeated by industrial trade

associations and politicians that oppose greenhouse gas regulations in support of efforts

to prohibit the use of the social cost of carbon. Environmental advocacy groups, on the

other hand, have argued that the social cost of carbon can be a useful tool for setting

climate policy, although they argue that current estimates are too low, because, for

example, they fail to adequately account for catastrophic outcomes (Environmental

Defense Fund et al. 2013; Weitzman 2009; Weitzman 2011; ). The social cost of carbon,

then, represents another example where the traditional interest group alignment

concerning cost-benefit analysis has been inverted by new political dynamics.

13

3. Instrument choice

Identifying the socially desirable level of environmental quality is only useful if

there are means to achieve that end. This section explores different types of government

interventions (i.e., policy instruments) that can be used to reduce pollution. It first

discusses normative questions and then examines instrument choices that have been made

in U.S. environmental policymaking. Section 4 will explore questions concerning the

level of government that is best suited to implement these instruments.

3.1 Normative Issues and Analysis

There are a wide range of policies that can be adopted to achieve environmental

goals, including labeling and disclosure requirements, technology standards, liability

regimes, effluent fees, and tradable emission allowances. These policies may differ along

a number of important dimensions, including the flexibility they provide for regulated

actors, their ease of enforcement, the information necessary to implement the policy, and

the incentives they provide for technological development on the part of private actors

(Goulder 2008). Selecting between instruments involves a sometimes complex inquiry

that is both conceptual as well as practical and context specific.

3.1.1 General Framework

For a given environmental goal, the cost-effective policy will be the one that

achieves that goal at the lowest total cost (Office of Management and Budget 2003.

While there is substantial controversy about appropriate environmental goals, the

aspiration of cost-effectiveness is widely shared. In general, cost-effective policies will

equalize marginal abatement costs across all pollution sources where reductions are

possible. When marginal abatement costs are not equal, lower-cost abatement

opportunities have not been fully implemented, implying that cost-effectiveness has not

been achieved.

An important qualification to the general principle that marginal costs be

equalized is the need to ensure compliance with environmental requirements.

Enforcement may be easier at some sources than others: point sources of pollution are

easier to monitor than non-point sources; inspection of a small number of large sources is

less costly than inspection of a large number of small sources; certain companies or

private individuals may be judgment proof, undermining the incentive effect of potential

penalties (Farmer 2007). For these reasons, and others, the aspiration of equalizing

marginal abatement costs may sometimes be subject to practical enforcement constraints.

Cost-effective policies will achieve environmental goals at the lowest aggregate costs,

including monitoring, inspection, and enforcement, meaning that some theoretical lower-

cost abatement opportunities may not be realized.

Dynamic effects may also provide reasons to depart from cost-effectiveness in the

short term. Environmental regulations often generate incentives for the development and

deployment of new technologies (Downing and White 1986; Ellerman et al. 2003;

14

Malueg 1989; Milliman and Prince 1989). Frequently, cost-effective instruments will

provide the correct incentive for efficient technological development: firms will make

marginal decisions to invest in new technologies or use existing technologies in ways that

minimize the present value of their compliance costs (Zerbe 1970; Downing and White

1986; Milliman and Prince 1989). Positive externalities associated with knowledge

spillovers, however, may result in underinvestment in technological development

(Katsoulacos and Xepapadeas 1996). In these cases, regulations designed to require firms

to overinvest in new technologies (relative to their private costs and benefits) may be

welfare maximizing (Jaffe et al. 2001). In the long run, this approach would be cost-

effective, but in the short term, lower-cost abatement opportunities may be foregone.

An additional element of instrument choice concerns the selection of regulatory

tools in the face of uncertainty. A social decision maker can be uncertain about either

environmental damages or abatement costs, or both. Weitzman (1974) shows that when a

regulator must either set a price (such as through effluent fee) or a quantity (such as

through a pollution cap) cost uncertainty can significantly affect this choice, with the

relative elasticity in damage and cost functions determining whether a price instrument is

more efficient that a quantity instrument. More recent research favors shows that non-

linear taxes, graded quantities, or hybrid tax-quantity instruments have advantages over

either flat prices or fixed quantities (Roberts and Spence 1976; Weitzman 1978; Kaplow

and Shavell 2002; Pizer 2002; Fell et al. 2012).

3.1.2 Market-mechanisms vs. command-and-control

A particularly important choice that is often presented in the environmental

context is between command-and-control style regulation and market-based approaches.

The typical example of a command-and-control environmental regulation is a design

standard that requires a specific pollution reduction technology to be adopted by all

regulated firms. An example of a market-based mechanism is a comprehensive,

economy-wide “cap-and-trade” system of tradable emission allowances. There are many

gradients along a command-to-market spectrum (Freeman and Kolstad, eds. 2006).

Performance-based standards, which set effluent limits but do not specify a particular

technology, are more market-like than a prescriptive design standard (EPA 2010b).

Flexibility can be added to a command-and-control standard by allowing firms to comply

with emissions requirement through offsets, in which new emissions must be

accompanied by equivalent reductions. A particular version of an offset mechanism is the

use of “bubbling” to treating a facility or firm as a single source. The use of bubbling

under the Clean Air Act new source performance standards provision was the substantive

question at the heart of Chevron v. NRDC,12

the case that sets out the contemporary

standard of judicial deference to agency statutory interpretations. Compliance flexibility

can also be enhanced through banking and/or borrowing, in which past or future

emissions reductions can “count” toward emissions limits. Environmental liability rules

(Viscusi and Zeckhauser 2011) and labeling and disclosure can be used to achieve

12 467 U.S. 837 (1984).

15

environmental goals (Thaler and Sunstein 2008) in place of, or as a supplement to, ex-

ante controls.

More market-like instruments are often favored on cost-effectiveness grounds

because they provide firms with flexibility in achieving emissions reductions and tend to

equalize marginal abatement costs across firms (Montgomery 1972; Baumol and Oates

1988; Tietenberg 1995). On the other hand, command-and-control mechanisms may be

easier to enforce (Grossman and Cole 1999) and they avoid problems that arise when

emissions are not spatially or temporally fungible (Sado et al. 2010).

3.1.3 Environmental transitions and grandfathering

As in all areas of policy change, environmental law creates issues of retroactivity

when private or public actions are undertaken under a prior regime that can, at least

potentially, be governed by the new regime (Fisch 1997). The issue of retroactivity can

be especially important in the environmental context because environmental standards

can affect decisions that are extremely widespread (such as car purchase decisions in the

context of fuel economy standards for automobiles) or touch on extremely high value

infrastructure investments (such as power plants affected by air quality standards). Many

environmental regimes adopt a bifurcated approach in response to the retroactivity

problem, treating new emissions sources differently than existing sources.

Three general justifications can be given for a bifurcated approach. First, existing

sources have a life-cycle of depreciation and technological obsolescence for reasons

unrelated to the environmental regime. Requiring expensive investments to upgrade

existing emissions sources that will soon be retired may be unjustified. Second, the

marginal cost of pollution control achieved by retrofitting existing sources may be much

higher than the same pollution reductions at new sources. Finally, public choice theory

may predict that existing sources may acts as an effective lobbying coalition against

economically justified environmental protections; grandfathering may be second-best

solution to overcome opposition to new controls (Revesz and Westfahl Kong 2011).

An important dynamic effect of the bifurcated approach is referred to as the “old

plant effect” (Ackerman and Hassler 1980). When additional costs are imposed on new

plants in the form of stringent environmental standards, the life of the existing plants is

extended and their replacement by new plants is delayed. The favorable treatment

extended to existing sources under the bifurcated approach increases their value, and

crowds out new sources based on more efficient technologies that would out-compete

existing sources in the absence of the environmental policy. The net result on emissions

from the imposition of an environmental standard when it is accompanied by an

exemption for existing sources is ambiguous (Nash and Revesz 2007). The old plant

effect is compounded by a public choice dynamic in which beneficiaries of a

grandfathering policy lobby to extend their favorable treatment for as long as possible

(Revesz and Westfahl Kong 2011).

16

Auction-based cap-and-trade or effluent fee systems represent an efficient

approach to allocating abatement costs and investments between new and existing

sources (Montero 2008). In cases where pure market-based mechanisms are infeasible,

bifurcated treatment may be justified, despite the old plant effect. The standard approach

to determining the optimal bifurcated treatment is through a two-step, sequential inquiry

(Shavell 2008). The first question is the optimal stringency for new sources; the second

question concerns the optimal transition rule for existing sources in light of the standards

for new sources. The sequential approach, however, does not account for the old plant

effect and therefore leads to better-than-optimal treatment of existing sources; jointly

optimizing the grandfathering rule and the new source standard leads to superior results

(Revesz and Westfahl Kong 2011).

3.2 Positive Issue and Analysis

Nearly a half-century after the United States embarked on major expansion in

federal environmental law, this natural experiment has generated some useful results that

can inform future policy design. In particular, experience with the sulfur dioxide trading

program established by the Clean Air Act Amendments of 1990 showed that market-

based mechanisms have great promise for achieving impressive emissions cuts and low

cost. Whether political leaders will profit from this experience is another question.

3.2.1 Experience with market-based mechanisms

As noted by Schmalensee and Stavins (2010), “market-based policies . . . were

innovations developed by conservatives in the [Ronald] Reagan, George H.W. Bush, and

George W. Bush administrations.” Given this pedigree, it is perhaps unsurprising that

many environmental groups opposed marketable permit schemes (Hahn and Hester

1989). The only major environmental organization that showed a strong interest in

developing market-based solutions to environmental problems, the Environmental

Defense Fund, was strongly criticized (Krupp 2008).

Developments in the late 1980s and early 1990s shifted this dynamic, holding out

the promise that a new political consensus in support of market mechanisms was

attainable. The passage of the Clean Air Act Amendments of 1990, which established a

national market for trading in sulfur dioxide emissions, is generally regarded as the first

large-scale collaboration across the political spectrum to support a marketable permit

scheme (Joskow and Schmalensee 1998). The legislation was adopted with wide

bipartisan support; only five Democratic and five Republican senators voted against the

bill.

The sulfur dioxide emissions trading program is widely viewed as a major success

(Chan et al. 2012). Impressive emissions reductions occurred at low costs (Ellerman et al.

1997). Use of the trading mechanisms allowed firms to deploy relatively low cost

alternatives such as fuel-shifting and other production process changes (Doucet and

Strauss 1994). The trading program also may have led to technological change that

17

would not have been induced through a design standard (Burtraw 1996; Ellerman and

Montero 1998; Bohi and Burtraw 1997; Keohane 2001). The program also resulted in

higher-than-expected benefits as the severe health consequences associated with

particulate matter (of which sulfur dioxide is a precursor) have become more clear

(Environmental Protection Agency 2011c).

The success of the U.S. sulfur dioxide program spurred major interest in tradable

allowance systems (Ayres 2000). Environmental issues for which tradable allowance

systems have been implemented include fisheries management and water quality (Shortle

and Horan 2006). Most important, a tradable allowance system is widely recognized as

the preferred approach to climate change among political leaders (Stavins 2008). The

European Emissions Trading System (ETS), created to fulfill obligations under the Kyoto

Protocol, is the most important existing greenhouse gas emissions allowances trading

system (IPCC 2007). Despite some continuing difficulties in maintaining price stability,

in part due to an excessively generous overall cap on emissions, the European ETS has

led to important emissions reductions and provided valuable lessons in market design

(Brown et al. 2012). Within the U.S., the political consensus around market-based

approaches to greenhouse gas reductions was nearly universal, with presidential

nominees of both major parties strongly supporting a cap-and-trade system and

environmentalists largely dropping their opposition to market mechanisms (Bernton

2008).

The promise of consensus, however, was short-lived. Comprehensive climate

change legislation was an early priority for the Obama administration. A bill to create a

comprehensive cap-and-trade regime was adopted by the Democratic-dominated House

of Representative without a single Republican vote.13

When the bill passed the House,

debate shifted to the Senate, where it took on a highly partisan tenor. Regulated industry

led a major lobbying and public relations push to oppose the legislation. Advocates and

politicians that opposed the bill cast it as a tax on energy that would hurt consumers

(Boehner 2009). The opposition campaign was ultimately successful, and the bill died in

the Senate without being called to a vote (Stromberg 2010). Opposition to cap-and-trade

became a centerpiece of the successful Republican effort to retake the House in 2010

(Good 2011). By the time of the 2012 presidential election, the issue had become a

political litmus test for conservatism in the Republican primary (Weigel 2011). There is

now a broadly shared view that legislation to create a market-based system to control

greenhouse gas emissions is not politically achievable in the current political

environment (Broder 2010). In the absence of legislation, EPA is moving forward with

regulations under the Clean Air Act (Executive Office of the President 2013). While the

statutory structure of the Act, especially section 115, may be broad enough to

accommodate a flexible, market-based approach (Chettiar and Schwartz 2009; Chang

2010), there is substantial legal uncertainty given the vagueness of the statutory language.

3.2.2 Labeling and disclosure

13 Office of the Clerk of the U.S. House of Representatives, “Final Vote Results for Roll Call 477,” available at http://clerk.house.gov/evs/2009/roll477.xml (June 26, 2009).

18

While the political consensus concerning market-based approaches to pollution

control was short lived, there has been renewed interest in labeling and disclosure as a

lower cost, and therefore more politically palatable, means of improving environmental

quality. In particular, Cass Sunstein’s tenure as the Administrator of the Office of

Information and Regulatory Affairs (OIRA) at the White House Office of Management

and Budget from 2009-12 was marked by efforts to improve the design of labeling and

disclosure regimes to maximize their effectiveness (Office of Information and Regulatory

Affairs [OIRA] 2010).

A new fuel economy label adopted in 2011 required all new automobiles to

“provide a clear statement about anticipated fuel savings (or costs) over a five-year

period” (Sunstein 2012). The label simplifies and clarifies fuel economy information

pertinent to consumer decision making, no longer “leav[ing] it to consumers to do the

arithmetic needed to figure out the net economic effects of fuel economy standards on

their budgets and lives” (Sunstein 2012; Environmental Protection Agency 2011d). The

design of the new label was foreshadowed in Sunstein and Thaler (2008), which argued

for display of multi-year estimates of fuel costs.

Similarly, in 2009 the Environmental Protection Agency promulgated a

mandatory greenhouse gas reporting rule14

which mimics disclosure programs like the

Toxic Release Inventory (TRI) required by the Emergency Planning and Community

Right to Know Act. The TRI, which requires reporting of both storage and release of

potentially hazardous chemicals, without requiring any additional action, has led to

significant reductions in toxic releases into the environment (Khanna et al. 1997). The

Greenhouse Gas Reporting Program (GGRP) will create a database covering 85–90

percent of total U.S. GHG emissions by requiring reporting from sources that “emit

25,000 metric tons or more of carbon dioxide equivalent per year in the United States,”

excluding the agricultural sector (Environmental Protection Agency 2013). The GGRP

will increase the quantity of greenhouse gas emission information, as well is its visibility

and salience, among both GHG sources and the general public (Office of Information and

Regulatory Affairs, 2010; Cohen and Viscusi 2012).

3.2.3 Social norms

There have also been recent efforts to use government institutions to influence

social norms in an environmentally-friendly direction, an approach that has received

attention as a “libertarian paternalist” mechanism to promote social goals (Sunstein and

Thaler 2008). One such example is an executive order15

that seeks to increase the

visibility and salience of energy costs within the federal government through methods

such as public scorecards and leadership awards (Sunstein 2011). As part of this effort,

the federal government has partnered with the private sector to set joint goals for energy

efficiency, leveraging its own efficiency improvements to spur improvements more

broadly (for example, U.S. Department of Energy 2013). These efforts expand on

14 40 C.F.R. pt. 98. 15 Executive Order 13,514, Federal Leadership in Environmental, Energy and Economic Performance, 74 Fed. Reg. 194 (Oct. 5, 2009).

19

voluntary partnerships programs that have existed in the environmental area for some

time (Borck and Coglianese 2009; Coglianese and Nash 2009).

3.2.4 Reducing the effect of grandfathering under the Clean Air Act

While the a bifurcated approach to new and existing sources is a common feature

of U.S. environmental law in recent years, the EPA has undertaken a number of

significant measures designed to limit the scope of grandfathering. For example, the

agency has engaged in a multi-year effort to establish a marketable permit scheme to

control interstate pollution that reduces favorable treatment for inefficient pre-Clean Air

Act sources. As discussed above, the most recent iteration of this effort, the Cross-State

Air Pollution Rule adopted in 2011, was struck down by the D.C. Circuit in 2012. Also in

2011, EPA adopted the Mercury and Air Toxics Rule (MATS), which limits the

emissions of a number of toxic air pollutants from both new and existing power plants.16

And, in 2013, President Obama indicated that in 2014, EPA will propose a rule limiting

the greenhouse gas emissions of existing power plants under section 111(d) of the Clean

Air Act.17

The combined effect of these approaches could significantly accelerate the

shutdown of existing plants that had been grandfathered for more than 40 years.

4. Jurisdictional allocation

The existence of externalities, transaction costs, and imperfect property rights

imply that an unregulated marketplace is unlikely to spontaneously provide efficient

levels of environmental quality—government intervention is necessary. Once that

preliminary observation has been made, there is a second question concerning how

governmental authority over environmental policy ought to be allocated between

different levels of government, from the local (municipalities) to the global (international

institutions). This question has important implications for the effectiveness and

legitimacy of environmental policy.

3.1 Normative Issues and Analysis

The allocation of policymaking responsibility across jurisdictions is an important

question in many legal domains, and one that has caught the attention of law and

economics scholars (Faure and Johnston 2009). Economics, in particular, can address the

role of incentives in determining whether a policymaking context is amenable or not to

local control. Environmental policymaking presents a number of specific issues that are

relevant to questions of jurisdictional allocation.

16 78 Fed. Reg. 79 (Apr. 24, 2013). 17 The White House, “Presidential Memorandum — Power Sector Carbon Pollution Standards,” available at http://www.whitehouse.gov/the-press-office/2013/06/25/presidential-memorandum-power-sector-carbon-pollution-standards (June 25, 2013).

20

3.1.1 General Framework

Several important inputs into environmental protection can vary by geographic

region, a fact that increases the desirability of locally tailored standards. These

geographically variable inputs include the marginal costs of pollution control, preferences

concerning the value of environmental quality (compared to other goods), and the level of

exposure to environmental risk (which is affected by population density, among other

factors) (Mendelsohn 1986; Nordhaus 1994). To the extent that information about

geographic variability is more likely to be held by local government officials, there is a

justification for a rebuttable presumption in favor of local control over environmental

policy (Revesz and Stavins 2007). At the same time, environmental protection is also an

area where, at least in some instances, the local-control presumption is rebutted by the

existence of inter-jurisdictional externalities (Revesz 1996).

Alternative justifications have been given for why national-level control over

environmental protection is desirable, but they have important weaknesses. Most

prominently, scholars have argued that environmental protection presents a “race-to-the-

bottom” problem in which inter-jurisdictional competition leads to inefficiently low

levels regulation (Esty 1996). Basic models have demonstrated, however, that rational,

self-regarding jurisdictions in a perfectly competitive market will arrive at efficient levels

of pollution control (Revesz 1992). If the assumption of rationality or perfect competition

is relaxed, jurisdictions may over- or under-regulate (Revesz 1997). Federal floors (which

are common in environmental law), then, are no better justified than federal ceilings

(which are relatively uncommon, although not unknown). In addition, to the extent that

environmental protection is federalized, jurisdictions may simply compete in other areas

more directly under local control and simply over- or under- provide some other public

good or service (Revesz 1997).

3.2 Positive Issues and Analysis

While economics provides a theoretically attractive framework for analyzing how

jurisdiction over environmental policymaking ought to be allocated, observed behavior

sharply diverges from its recommendations.

3.2.1 Jurisdictional mismatch

The allocation of authority between the national government and the states in U.S.

environmental law is not tailored to a jurisdiction-externality justification for federal

authority. There are many federal statutes that address environmental problems that do

not have any inter-jurisdictional externalities. These include federal programs to

remediate hazardous waste sites (the “Superfund”),18

and set limits on the allowable level

of contaminants in drinking water.19

Furthermore, the Clean Air Act and Clean Water

Act, which do address pollution sources with the potential to generate important interstate

18 Comprehensive Environmental Response, Compensation, and Liability Act, 42 U.S.C. § 103 et seq. (2006). 19 Safe Water Drinking Act, 42 U.S.C. § 300f et seq. (2006).

21

externalities, are generally focused on local, rather than inter-state, pollution. The core of

the Clean Air Act consists of federally prescribed air quality standards designed and

implemented at the local level.20

Indeed, facilities can meet these air quality standards by

exporting more pollution across state lines (Revesz 1996). Plant level emissions standards

are not oriented toward facilities with important inter-state consequences, instead

covering all sources by pollution category and vintage.21

The sections of the Clean Air

Act that are specifically targeted toward interstate pollution have been devilishly difficult

for the EPA to implement, with multiple attempts being struck down by the D.C. Circuit

Court of Appeals.22

The Clean Water Act fares little better, with most of its emphasis

placed on pollution sources that have only intrastate effects (Stewart 1982).

3.2.2 Lack of a global agreement on climate change

If governance power should be allocated to the minimally extensive jurisdiction

that internalizes any relevant externality, in the context of climate change, that

jurisdiction is the entire globe. Because greenhouse gas emissions in any country lead to

climate change risks in all countries, a global approach to greenhouse gas limits is well-

justified.

There has been a substantial effort through the United Nations Framework

Convention on Climate Change (UNFCCC) to negotiate a set of meaningful mandatory

limits on emissions that would apply to all countries. While the UNFCCC process has

had some successes, it has faced extremely serious stumbling blocks in recent years. In

particular, the failure to negotiate a successor agreement to the Kyoto Protocol at the 15th

Conference of the Parties meeting in Copenhagen, Denmark was seen as a major setback

for the IPCCC process (Vidal 2009). Renewed efforts have focused on the 21th

Conference of the Parties meeting in Paris, France, to occur in 2015.

The lack of progress at the international level has resulted in the devolution of initiative

for climate policy to the regional or domestic level. The European ETS remains the most

robust international emissions trading program and serves as the primary vehicle for

emissions reductions within European countries (European Commission 2013). Domestic

efforts to curb greenhouse gas controls have been forwarded successfully in some

countries, but met with stiff resistance in countries like Australia (Plumer 2013).

3.2.3 State innovation

Within the United States, the lack of national emissions limits, and especially the

lack of climate legislation, has resulted in a further devolution to the state level (Carlarne

20 See Clean Air Act §§ 108–109, 42 U.S.C. §§ 7408–7409 (2006). 21 § 111, 42 U.S.C. § 7411 (2006). 22 See North Carolina v. EPA, 531 F.3d 896 (D.C. Cir) (striking down the Clean Air Interstate Rule); EME Homer City Generation, L.P. v. EPA, 696 F.3d 7 (D.C. Cir. 2012) (striking down the Cross State Air Pollution Rule).

22

2008). While this state of affairs may not be optimal from the perspective of economic

theory, it is nonetheless a political reality.

U.S. states have adopted three basic approaches. Early efforts tended to involve

command-and-control regulation. In 1997, Oregon enacted the first legislation in the

United States addressed at limiting greenhouse gas emissions, setting a standard for

carbon dioxide emissions from the state’s natural gas electric plants (Environmental

Defense Fund 2012). In 2001, Massachusetts enacted carbon dioxide regulations for all

power plants, as part of a comprehensive bill aiming to cut pollution from the electricity

sector (Daley 2001). In 2002 California enacted legislation that required the state

regulatory agency to “adopt regulations that achieve the maximum feasible and cost-

effective reduction of greenhouse gas emissions from motor vehicles.”23

The California effort was especially significant because the state plays a special

role under the Clean Air Act (Carlson 2009). Section 209 of the Act allows California to

request a waiver from the EPA to set more stringent mobile source standards than the

federal standards.24

Other states can then follow suit, and depart from the federal

standard, if they choose. When California issued regulations in 2004 establishing

greenhouse gas limits for motor vehicles, it set in motion a political chain of events that

ultimately led to a “car deal” negotiated between major automobile manufacturers, the

federal government, and other stakeholders in support of a national greenhouse gas

standard for new automobiles under the Clean Air Act (Freeman 2011).

Subsidies to encourage greater reliance on renewable sources of electricity are a

second approach adopted by many states. The most common approach, referred to as

“renewable portfolio standards,” is a mandate that a target share of the state’s energy

supply be generated by renewable sources (Davies 2012). For example, California’s

standard requires suppliers to obtain 20% of their energy from renewable sources, with

this proportion rising to 33% by 2020 (Farber 2008). Clean energy subsidies face

important challenges, including the need to define “renewable” in a manner that

accurately captures environmental benefits (Duane 2010), and likely result in relatively

high cost emission reductions (OECD 2013).

Finally, states have experimented with a cap-and-trade approach to greenhouse

gas emissions limits. The Regional Greenhouse Gas Initiative (RGGI) is one such effort.

RGGI is a collaboration of northeastern states that have signed a joint memorandum of

understanding (MOU) setting out each state’s share of a regional carbon dioxide cap,

adopted legislation or regulation approving that MOU, and, beginning in 2008,

implemented a function auctioning and trading process (Duane 2010). While one state

from the initial group of ten (New Jersey) has withdrawn, the RGGI regime has remained

relatively stable and has raised over $1 billion in the first four years of operation for

participating states (Rabe 2009).

23 California Assembly Bill 1493 (July 22, 2002), available at http://www.leginfo.ca.gov/pub/01-02/bill/asm/ab_1451-1500/ab_1493_bill_20020722_chaptered.pdf. 24 Clean Air Act § 209, 42 U.S.C. § 7543 (2006).

23

In 2006 California passed the California Global Warming Solutions Act,

commonly known as AB 32.25

That legislation launched a multi-step process to reduce

greenhouse gas emissions in the state to 1990 levels by 2020. The centerpiece of the

regulatory approach implementing AB 32 is a statewide cap-and-trade program.

California has not, however, adopted a purely market-based approach, having augmented

its cap-and-trade approach with a range of additional policies, including plant specific

performance standards, energy efficiency requirements, and a renewable portfolio

standard. Many of these additional measures are likely to lead to increased costs without

obvious climate benefit (Carlson 2013).

5. Conclusion

Since the publication of Revesz and Stavins (2007), there have been some

significant normative advances in the area of environmental law and economics. For

example, the emergence of climate change as the area of central concern for

environmental regulation has brought a great deal of attention to the question of how to

discount benefits that accrue into the far future and primarily affect individuals not yet

born. Also, the rise of behavioral law and economics has created a shift away from

exclusive reliance on neoclassical models.

But the most significant changes have been on the positive side. In particular, the

traditional alignment of interest groups has come close to experiencing an about-face.

Conservative, anti-regulatory groups traditionally favored cost-benefit analysis, market-

based instruments, and decentralization. Progressive, pro-regulatory groups traditionally

opposed these approaches. In recent years, however, the tables have often been turned.

These shifts suggest that commitment to principles is secondary to commitment to

substantive regulatory outcomes, with groups of both sides of the spectrum availing

themselves of whatever argument will better promote their preferences concerning the

stringency of regulation.

25 Global Warming Solutions Act, California Assembly Bill 32 (Sept. 27, 2006), available at http://www.leginfo.ca.gov/pub/05-06/bill/asm/ab_0001-0050/ab_32_bill_20060927_chaptered.pdf.

24

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