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date: 17 November 2017

Payments versus Direct Controls for Environmental Externalities in Agriculture

Summary and Keywords

The production of food, fiber, and fuel often results in negative externalities due to impacts on soil, water, air, or habitat. There are two broad ways to incentivize farmers to alter their land use or management practices on that land to benefit the environment: (1) provide payments to farmers who adopt environmentally beneficial actions and (2) introduce direct controls or regulations that require farmers to undertake certain actions, backed up with penalties for noncompliance. Both the provision of payments for environmentally beneficial management practices (BMPs) and a regulatory requirement for use of a BMP alter the incentives faced by farmers, but they do so in different ways, with different implications and consequences for farmers, for the policy, for politics, and consequently for the environment. These two incentive-based mechanisms are recommended where the private incentives conflict with the public interest, and only where the private incentives are not so strong as to outweigh the public benefits. The biggest differences between them probably relate to equity/distributional outcomes and politics rather than efficiency. Governments often seem to prefer to employ beneficiary-pays mechanisms in cases where they seek to alter farmers’ existing practices, and polluter-pays mechanisms when they seek to prevent farmers from changing from their current practices to something worse for the environment. The digital revolution has the potential to help farmers produce more food on less land and with fewer inputs. In addition to reducing input levels and identifying unprofitable management zones to set aside, the technology could also alter the transaction costs of the policy options.

Keywords: agri-environmental policy, payment schemes, direct controls, economic incentives, regulation, externalities, property rights, transaction costs

Introduction

The production of food, fiber, and fuel often results in negative externalities due to impacts on soil, water, air, or habitat. The extent of the externality varies both spatially and temporally, depending on the production techniques used and the biophysical context. The Organisation for Economic Co-operation and Development (OECD) has developed agri-environmental indicators to measure the degree of damage to ecosystems caused by farming practices (OECD, 2015). Across OECD countries, on the one hand, the most severe negative externalities from agriculture are those associated with a decline in biodiversity and an increasing risk of water contamination from excess nutrients. On the other hand, farming activities can also generate positive externalities such as through the provision of rural amenity the value of which depends on the alternative use of the agricultural land.

There are two broad ways to incentivize farmers to alter their land use or management practices on that land to benefit the environment: (1) provide payments to farmers who adopt environmentally beneficial actions, and (2) introduce direct controls or regulations that require farmers to undertake certain actions, backed up with penalties for noncompliance. The first approach is the most commonly used agri-environmental policy across developed countries. One example is the Conservation Reserve Program, which pays farmers in the United States to set aside environmentally sensitive cropland from production (Claassen & Ribaudo, 2016). Examples of direct regulation (the second approach) include the banning of certain pesticides (Davis, 2014) and the requirement for buffer strips around watercourses to limit nutrient loading (Sieber et al., 2010). Note that payments (subsidies or taxes) and regulations are categorized here as incentive systems because both encourage changes in management. Some associate incentives only with payments that shape voluntary behavior and not with regulations that impose involuntary constraints on behavior. However, both approaches are considered as incentive systems since rewards or penalties are created for the decision maker to alter the practices used.

Both the provision of payments for environmentally beneficial management practices (BMPs) and a regulatory requirement for use of a BMP alter the incentives faced by farmers, but they do so in different ways, with different implications and consequences for farmers, for the policy, for politics, and consequently for the environment. Both are incentive mechanisms that are typically design-based (focused on BMPs), but the beneficiary pays in the first approach, while the farmer or polluter pays in the case of direct controls and/or financial penalties (the second approach). In assessing the implications for the design and implementation of agri-environmental policies, the following questions are addressed here:

  • Under what circumstances is it appropriate to implement an incentive-based mechanism (of either type) to address environmental externalities from agriculture (as opposed to alternatives such as communication, persuasion, or R&D)?

  • Where an incentive-based mechanism is used, what are the implications and consequences of each option (payments versus control) for farmers, for the policy, for politics, and consequently for efficient and equitable protection of the environment?

  • Under what conditions is each policy option likely to be more effective and efficient?

Environmental Externalities in Agriculture

Types of Externalities

Before proceeding with the evaluation of alternative policy mechanisms, it is helpful to understand the types of externalities generated by agricultural activities, as the effectiveness of any policy depends on the nature of the problem at hand. The emphases of agri-environmental policy are different in different countries, depending on local conditions, preferences, and history. OECD (2015) provides an overview of priorities for five OECD countries (Table 1). In their assessment, the most important externality-related issue across these five countries, on average, is protection of biodiversity, followed by water pollution and agricultural landscapes. Even for these issues, there is variation in their assessed importance between countries, with water pollution judged to be of low policy importance in Japan and agricultural landscapes scored as low importance in the United States and of no policy relevance in Australia. A discussion of each of the externality types follows.

Table 1. Agri-environmental public goods targeted in several OECD countries (0 = not targeted; X = low importance; XX = medium importance; XXX = high importance).

Externality

Australia

Japan

Netherlands

United Kingdom

United States of America

Soil protection and quality

XX

X

X

XX

XX

Water quality

XX

X

XX

XXX

XXX

Air quality

X

X

XX

XX

X

Climate change—emissions

X

XX

XX

XX

0

Climate change—sequestration

X

X

XX

XX

0

Biodiversity and habitat

XXX

XX

XXX

XXX

XXX

Agricultural landscapes

0

XX

XXX

XXX

X

Source: Adapted from OECD (2015).

Soil Protection and Quality

Soil quality attributes often affect the farm-level economics of production, as well as having off farm impacts. For example, the practice of zero tillage plus retention of crop residues enhances soil structure and increases yields (Gray et al., 1996; Fulton, 2010), while also reducing soil erosion that would otherwise contribute to dust storms or sedimentation of dams (Giller et al., 2009). Soil salinity in Australia reduces on-farm productivity of crops and pastures, but may also have impacts on water bodies, native vegetation, and nearby built infrastructure such as roads (Pannell, 2001).

Water Quality

Off-farm water quality is not only affected by soil erosion and salinity but also by nutrient pollution originating from organic and inorganic fertilizers. Notable examples include algal blooms in Chesapeake Bay (Hagy et al., 2004) and Lake Erie (International Joint Commission, 2014), the hypoxic zone in the Gulf of Mexico (Rabotyagov et al., 2014), and reduced water quality affecting the health of the Great Barrier Reef (Wooldridge, 2009). A further source of water pollution in some cases is pesticides and herbicides washing into rivers in surface runoff from croplands (Reichenberger et al., 2007), affecting ecology and human health.

Air Quality

Impacts on air quality from agriculture include odors, ammonia, and dust. Although direct physical human health damages are possible, the major adverse impacts from odors associated with waste from livestock farming are nuisance-related. Complaints surrounding intensive livestock operations and associated odors tend to increase with the education and income levels of the nonfarm population in the region (Weersink & Raymond, 2007). Ammonia emitted from livestock waste affects odor but also reacts in the atmosphere to form particulates that can damage plant ecosystems (Bitman et al., 2014).

Climate Change (Emissions and Sequestration)

Agricultural emissions of CO2 equivalents constitute around 8% of global emissions (OECD, 2015). As well as emission sources common to other sectors, such as use of fossil fuels for energy, agriculture additionally emits large amounts of ammonia from livestock (Jayasundara et al., 2016) and N2O from fertilizer (Reay et al., 2012). For example, approximately 10% of Canada’s total greenhouse gas emissions are from crop and livestock production, with 25% of this total in the form of nitrous oxide associated with fertilizer use (ECCC, 2016). These are particularly potent greenhouse gases with high warming potential. Agriculture, however, also provides potential external benefits through sequestration of CO2 in soils or wood (Lal, 2004).

Biodiversity and Habitat

In long-settled parts of the world, species have coevolved with agriculture, resulting in ecosystems that are dependent on the continuation of traditional agricultural practices (Swinton et al., 2007). Maintaining these ecosystems on farms provides external benefits to nearby residents and others who care about them. In more recently settled areas, such as Australia, the focus of policy for biodiversity and habitat on farms has been on maintenance and protection of remnant native vegetation that was not cleared when the farms were established (Polyakov et al., 2015).

Agricultural Landscapes

In some countries, agricultural landscapes have evolved over many centuries, and their appearance is positively appreciated by nonfarmers resident in the region. For example, an aim of the Common Agricultural Policy in Europe is to maintain relatively traditional agricultural farming systems because their appearance is appreciated by many people (Lefebvre et al., 2015, van Zanten et al., 2014). In certain countries, notably the United Kingdom, agricultural landscapes are appreciated as a place for recreation (Dwyer, 2014).

Complications in Agri-Environmental Policy Design

The six agricultural externalities from Table 1 (combining the two climate change ones) all have public-good characteristics: nonexcludability and/or nonrivalry (Bergstrom & Randall, 2016). For example, the aesthetic benefits of agricultural landscapes are inherently both nonexcludable and nonrival. Recreational benefits such as walking on agricultural land are nonexcludable by policy choice in some countries (e.g., the United Kingdom) but not others (e.g., Australia). Water pollution from agriculture usually comes from nonpoint sources and so is nonexcludable, due to the difficulty of detecting the source.

Many nonagricultural externalities involve a moderate number of polluting firms with emissions that are relatively easy to monitor and measure, making the design and implementation of policy relatively easy. In contrast, agricultural externalities tend to be generated by many small, heterogeneous firms through complex biological processes subject to stochastic events. As a result of the interaction of these components, there is no single policy option consistently used to improve the environmental performance of agriculture (Weersink et al., 1998).

The large number of small potential contributors to the environmental problem makes it costly to measure emissions from each individual operation. In addition to making it more difficult to separate damages across farms, the large number of potential polluters decreases the likelihood of those farms working together to solve the problem.

Assigning liability is further complicated by the complexity of the production and environmental process, further enhancing the difficulty of measuring the emissions from each individual farm. The link between the adoption of beneficial management practices and external impacts at a certain location is often highly uncertain and difficult to quantify. This makes it difficult, even for relevant specialists, to determine whether the use of a particular BMP should be recommended to particular farmers in a particular place, or to assign liability for impacts to particular farmers. Even if a BMP is itself easily observable, (e.g., tillage methods or crop choice), the external impacts from a given location employing those BMPs depends on many other factors that may be more difficult to observe (e.g., timing, groundwater depth).

Spatial, temporal, and technological heterogeneity in the generation of external impacts from agricultural production influences the distributional consequences of the policy and thus its social and political acceptability. The levels of pollutants generated and their impacts on environmental health will depend not only on the farming system but also on geographic factors such as soil type, slope, and weather. Similarly, the time lag between the actions of a producer and damages resulting from those actions will depend on the type of externality and the location. For example, the depth of the water table and the soil type influence the time it takes for excess nitrogen to push up nitrate levels in the groundwater. The differences in soil type can also influence the abatement cost of implementing a BMP such as conservation tillage; farmers on sandy soil are likely to adopt conservation tillage without regulation. The greater the spatial, temporal, and technological heterogeneity, the greater the efficiency of targeted policy options.

Incentive-Based Mechanisms Versus Other Mechanisms

Given the range of environmental externalities associated with agriculture and the complications associated with the generation of these externalities, there is not a single best policy instrument, such as introducing a Piguovian tax on emissions equal to their marginal external cost, as suggested by many textbooks (Goulder & Parry, 2008). Instead, policymakers must decide who to target, what should be targeted, and which mechanism to use (Shortle & Horan, 2001).

Pannell (2008) categorized policy mechanisms for managing externalities into the following: (1) positive incentives (financial or regulatory instruments to encourage change) (2) negative incentives (financial or regulatory instruments to inhibit change), (3) extension (technology transfer, education, communication, demonstrations, support for community network), and (4) technology change (development of improved land management options, such as through strategic R&D, participatory R&D with landholders, or provision of infrastructure to support a new management option). A fifth option available to governments is) no action, which can be appropriate if the cost of achieving a desired change is so large that it outweighs the resulting benefits, or if the benefits are expected to occur even without government intervention.

When choosing a policy mechanism to achieve the target, policymakers should balance the value of reducing the externality to the direct economic costs incurred in achieving that benefit (Richards, 2000). Pannell (2008) developed the Public:Private Benefits Framework, a simple tool that helps to identify which of these policy instrument categories is most suitable to deal with externalities from agriculture, based on the levels of public and private net benefits that are likely to result from the land-use change. Here it is used to understand the contexts in which an incentive-based mechanism of any type is appropriate.

The consideration of public and private net benefits noted by Richards (2000) is illustrated in Figure 1 developed by Pannell (2008). Public net benefits consist of the external benefits and costs arising from a change in land use or land management and are measured along the vertical axis. The horizontal axis represents private net benefits to the farmer of the management system. The zero–zero point in the center of Figure 1 represents the current practice used by producers no matter how good or bad that practice is with regard to public net benefits.

Payments versus Direct Controls for Environmental Externalities in AgricultureClick to view larger

Figure 1. Recommended efficient policy mechanisms based on a simple set of rules.

Now, for any given project, a user of the framework estimates the levels of public net benefits and private net benefits from the project, relative to current practice, and plots the result on the graph. Depending on the location of the project on the graph, the appropriate policy response is indicated. The relevant questions addressed by the framework are (1) whether it is possible and worthwhile to do better than current practice, (2) whether it is worthwhile stopping or discouraging landholders from switching from their current practice to something more environmentally damaging, and (3) if so, how? These remain relevant questions whether the current practice is highly damaging or highly beneficial to the environment.

The Public:Private Benefits Framework aims is to identify which category of policy mechanisms is likely to be suitable for each potential means to address an externality from agriculture. The framework is based on the following set of rules for selecting a policy mechanism that are captured in Figure 1.

  1. 1. Do not use positive incentives to change land-use/management unless public net benefits of change are positive.

  2. 2. Do not use positive incentives if landholders would adopt land-use/management changes without those incentives.

  3. 3. Do not use positive incentives if private net costs outweigh public net benefits.

  4. 4. Do not use extension unless the change being advocated would generate positive private net benefits. In other words, the practice should be sufficiently attractive to landholders for it to be “adoptable” once the extension program ceases.

  5. 5. Do not use extension where a change would generate negative net public benefits. Note that Rules 4 and 5 are referring to cases where extension is used as the main tool to achieve land-use change. Extension could also be used to support any of the other policy mechanisms, playing a supporting role, rather than being the main tool.

  6. 6. If private net benefits are negative (but not overly negative), consider technology development to create improved (environmentally beneficial) land management options that can be made adoptable with or without positive incentives (Pannell, 2009).

  7. 7. If private net benefits outweigh public net costs, the land-use changes could be accepted if they occur, implying no action, or they could be penalized at an appropriate level but not prohibited. The latter approach uses a pricing mechanism to force landholders to consider the negative consequences of their actions. This allows them to weigh whether their benefits exceed those negative consequences, thus making prohibition unnecessary.

  8. 8. If public net costs outweigh private net benefits, use negative incentives.

  9. 9. If public net benefits and private net benefits are both negative, no action is necessary. Adverse practices are unlikely to be adopted.

  10. 10. In all cases, the suggested action needs to be weighed against a strategy of no action.

This simple framework can be made more sophisticated in various ways. When estimating net benefits, if there are time lags until costs or benefits are realized, these should be discounted using standard discounting methods. Thus, the public and private net benefits that are graphed would be “present values.” Allowing for time lags until adoption, learning costs involved in land-use change, the fact that extension reduces but does not eliminate time lags to adoption, the transaction costs of extension, and requiring higher levels of selectivity (a higher benefit: cost ratio) than just covering costs would alter both the public and private net benefits illustrated in Figure 1. Figure 2 allows for these complexities and requires a benefit:cost ratio of at least 2. A smaller number of projects would qualify for incentives or extension in this more targeted approach, relative to Figure 1.

Payments versus Direct Controls for Environmental Externalities in AgricultureClick to view larger

Figure 2. Efficient policy mechanisms for encouraging land use on private land, refined to account for various complexities described in the text.

In broad terms, the framework advocates the use of:

  • positive incentives (to encourage change) if the public net benefits of land-use change are high and the private net benefits are not too negative;

  • negative incentives (to discourage change) if private net benefits are less than public net costs;

  • extension if the public net benefits of land-use change are high and the private net benefits are moderate;

  • no action if private net benefits are positive and public net benefits are not sufficiently high;

  • no action if private net benefits are greater than public net costs;

  • no action if public net benefits and private net benefits are both negative;

  • technology development if private net benefits are low-to-moderately negative and public net benefits are positive (Pannell, 2009).

The framework makes clear that incentive-based mechanisms are not the most appropriate response in all cases. They are recommended to encourage favorable changes and discourage unfavorable changes but only in cases where the private incentives conflict with the public interest, and only where the private incentives do not outweigh the public benefits.

Types of Incentive-Based Mechanisms

The Public:Private Benefits Framework does not specify which type of incentive mechanism is recommended—only whether any type of incentive mechanism is appropriate. Within the broad group of incentive-based mechanisms, they can be categorized as either polluter-pays mechanisms or beneficiary-pays mechanisms (including cost-sharing approaches where beneficiaries do not cover the full costs). Here we expand on the mechanism options within these categories.

Polluter-Pays Mechanisms

These mechanisms consist of regulatory controls backed up by penalties for noncompliance or of taxes on externalities or associated inputs. They influence incentives because (1) the farmer is faced with an increased cost, or an increased risk of a cost, for each unit of the externality generated, and (2) they create signals showing what behavior is socially acceptable. Within this category, we identify four subcategories.

Performance-Based Standards or Penalties

Direct controls can be imposed on the acceptable limits of resource quality (ambient standard) or on the effluent emitted from the polluting source (emission standard). Such controls, particularly emission standards, are common in sectors with relatively few polluters generating measurable effluent. Firms facing these direct controls typically file evidence on their pollutant levels to the government, and fines, penalties, or taxes are imposed if the emissions exceed the allowable standards. While farms are accustomed to meeting standards on the quantity and quality of their produce (good output), performance-based environmental standards are rare in agriculture owing to the diffuse nature of most externalities, making it costly to measure emissions.

The design-based standards need not be regulated directly but can serve as the basis for a certification program that producers can voluntarily join. Environmental stewardship certification schemes confer benefits either as legal protection to certified farmers (e.g., organic certification) or as marketing signals (e.g., Michigan Agricultural Environmental Assurance Program; Vollmer-Sanders et al., 2011). Certification efforts such Field to Market and the 4 Rs programs provide evidence to government regulators or consumers willing to pay a premium for the environmental benefits that the stipulated design-based practices are being followed (Waldman & Kerr, 2014; Garnache et al., 2016).

An alternative is to estimate emissions through a biophysical simulation model that accounts for site-specific characteristics and the known impacts of management practices. The approach has been used in the Netherlands to tax surplus nitrogen and phosphorus from livestock farms (Breembroek et al., 1996). Nutrient surplus is estimated using information on the quantity and quality of feed and manure along with livestock numbers. The costliness of the approach is moderated to some extent by the ability to use farm-specific information collected for other purposes, and it is judged to be worth the public investment because of the extent and importance of the nutrient loading problem in water bodies in the country.

Although not falling as Incorporate environmental stewardship certification schemes. Public environmental certification programs typically confer benefits either as legal protections to certified farmers (e.g., USDA Organic certification) or as marketing signals (e.g., Michigan Agricultural Environmental Assurance Program; Vollmer-Sanders et al., 2011). But certification schemes seem to growing in the United States, especially in the private sector (e.g., Field to Market, 4 Rs programs). See Waldman and Kerr (2014) and Garnache et al. (2016).

Design-Based Standards or Penalties

Given the difficulty of determining emissions from individual farm operations, standards tend to be imposed on observable practices related to emissions. Initially, restrictions were focused on the use of pesticides, and such controls continue into the 21st century, as evidenced by the ban of certain neonicotinoid pesticides in the European Union and Canada owing to concerns about their effect on pollinator populations. Restrictions are now applied to a much broader range of agricultural inputs and practices. For example, crop producers in some jurisdictions cannot purchase pesticides without having attended safety and application method workshops, while livestock producers cannot apply manure to fields during the winter time so as to reduce the risk of nutrient runoff during spring melt. The issuance of a building permit for a new barn often involves restrictions on the location and sizing of the new facility, such as minimum separation distance between waterways and manure storage facilities (Weersink & Eveland, 2006).

The other way to create a polluter-pays incentive is through a tax on inputs, such as pesticides and fertilizer, correlated with the negative externality. The purpose of the tax is to raise the price of the polluting inputs and incentivize farmers to reduce its use and therefore the associated externality. The effectiveness of such a tax depends on the price elasticity of demand for the inputs (Burrell, 2012). In the case of pesticides, Bocker and Finger (2017) find the elasticities for pesticide use to be generally very inelastic, but the response in individual situations depends on factors such as crop type, time horizon, and pesticide type. Where that elasticity is low, an input tax may still be used to raise tax revenues to fund other programs to either alter farmer behavior or support R&D into more environmentally friendly but also more profitable BMPs (Zilberman & Millock, 1997).

Liability Laws/Performance Bonds

Liability laws allow the victims of environmental damages to file a suit against the polluters seeking compensation for the damages inflicted. Litigation has been used to deal with agricultural externalities in relatively rare situations such as an infrequent pollution event (e.g., a manure spill), in which the damages are significant, the cause and effect are known, and few parties are involved (Weersink et al., 1998).

Most regulatory approaches require that a fine be paid after damages have been incurred. An alternative mechanism is to require firms to post a bond that would be forfeited in part or in full if a pollution event occurs owing to inadequate control. An example of a performance bond in agriculture is the permit and damage deposit required by large, custom handlers of livestock waste. As with liability laws, performance bonds are effective in situations with easily observable links between the actions of a firm and the environmental externality.

Tradeable Permit Schemes

Coase (1960) highlighted that a lack of well-defined property rights can be a cause of externalties persisting at inefficient levels. If rights to pollute or to be free of pollution were clearly owned and tradeable, an efficient outcome could be achieved through bargaining and negotiation, without the need for the government to set regulatory standards or make payments. Governments can foster this approach by establishing a tradable pollution permit scheme.

The market for such permits allows firms with high abatement costs to purchase pollution permits from firms with lower costs, thereby resulting in a reduction in total abatement costs as compared to a scenario in which the pollution abatement of all firms is determined by regulatory standards. The success of a tradable pollution market for the reduction of sulfur dioxide in the United States (Schmalensee & Stavins, 2013) spurred interest in its use to deal with other externalities, including within the agricultural sector. Nutrient-based trading programs have been attempted in many countries and typically involve trades between point sources, such as municipal wastewater treatment plants, and farmers, who may have lower abatement costs (Fisher-Vanden & Olmstead, 2013). However, the success of such water-quality trading programs has been mixed, with some success where the major contributors to the problem are point sources that face significant emission reduction targets (King, 2005) but failures in other cases (Shortle, 2013).

In some jurisdictions, a tradable permit market exists for carbon dioxide equivalents. So far, agriculture has tended not to be included in these schemes as a direct source of greenhouse gas (GHG) emissions, partly due to the transaction costs of measuring and trading small amounts of emissions from many farms. In some cases, agriculture benefits by being granted the right to sell permits to the market where they sequester carbon through conservation tillage or the planting of perennials. However, the transient nature of carbon in soil organic matter makes it difficult to measure the extent of the carbon sequestered from these practices.

Beneficiary-Pays Mechanisms

Performance-Based Payments

Payments are widely used to incentivize farms to act in an environmentally beneficial manner. Performance-based payments, as is true of performance-based standards discussed earlier, require the ability to accurately measure emissions at relatively low cost. In practice, both performance-based approaches are rarely used to deal with agricultural externalities because of the costliness and difficulty of measuring environmental outcomes at the individual farm level for most pollutants.

Design-Based Payments

One of the most common incentive mechanisms used to address agricultural externalities is to offer incentive payments to encourage the use of BMPs or the production of certain outputs that improve environmental quality. A recent example is the Agricultural Greenhouse Gases Program (AGGP), which is a $27 million, five-year (2016–2021) Canadian program that offers cost-share grants to enhance the adoption of BMPs that will mitigate agricultural GHG emissions (AAFC, 2016). Similarly, the Great Lakes Agricultural Stewardship Initiative (GLASI) is a 2016 initiative to improve soil health and water quality through promoting the adoption of BMPs (OSCIA, 2016).

If such payments are large enough, they may compensate farmers for the full opportunity costs of adopting environmentally beneficial practices that are unprofitable. One way to ensure that the payments are large enough is to disperse funds through a reverse auction, where payments are made according to bids received from farmers, with funds targeted to those bids that provide the best value for money (environmental benefits per dollar spent; Boxall et al., 2017). This approach has been used, for example, in the Conservation Reserve Program (Claassen & Ribaudo, 2016) and the BushTender scheme in Australia (Stoneham et al., 2003). A downside noted by Palm-Forster et al. (2016) is that the theoretical cost effectiveness of reverse auctions is undermined by strategic bidding and the transaction costs of participation.

Most of the design-based programs used in agri-environmental policy, however, are universal cost-share arrangements; a set of eligible BMPs is eligible for funding support up to a certain percentage for any farmers within a wide region. Their universality (i.e., untargeted nature) and indirect linkage to the externality reduces the effectiveness of design-based payment schemes. Farmers with the greatest contribution to the externality may not apply if they cannot manage the BMP or if the financial incentive is too small. In contrast, funding may be granted to farms even though adoption was already planned (additionality problem) or to farms not contributing to the externality because the subsidy now makes the practice profitable (Weersink & Wossink, 2005).

In some cases, farmers are offered small temporary payments for using BMPs (e.g., most payment-based programs in Australia). This approach can be useful in acceleratingthe trial and adoption of a new practice that is judged to have positive private net benefits (as well as positive public net benefits) but has not yet been widely adopted. Duke et al. (2013) offer suggestions for improving the cost effectiveness of design-based conservation programs.

Rather than payments to encourage the use of BMPs to reduce a negative externality, payments can be used to encourage the uptake of practices that generate positive externalities, such as attractive agricultural landscapes. Baylis et al. (2008) note that Europeans are willing to pay for such positive externalities; these payments are built into the Common Agricultural Policy (CAP).

Additional Considerations in the Choice between Incentive Mechanisms

Who Pays?

The “beneficiary-pays principle,” closely related to the “user-pays principle,” states that the beneficiary of a good or service should bear the costs of its provision. The “polluter-pays principle” moves the financial burden onto those who are creating the environmental problem. Polluter pays was recommended as the default position by the OECD (2010). In reality, however, neither of these ideas is actually a principle in the sense of a scientific or mathematical principle: a truth that can be proved based on some other truths. Indeed, polluter-pays and beneficiary-pays are usually in direct contradiction, so they cannot both have the status of truths. Rather, they are best thought of as rules of thumb that might or might not be considered fair. The role of deciding who should bear the costs appropriately falls to governments.

In practice, governments often seem to prefer to employ beneficiary-pays mechanisms in cases where they seek to alter farmers’ existing practices, and they prefer polluter-pays mechanisms when they seek to prevent farmers from changing their current practices to something worse for the environment. In other words, the status quo commonly defines de facto property rights. To the extent that this generalization is true, it may reflect community judgments about what is fair and equitable, which probably affect the political consequences of the alternatives.

It is important to point out the practical limitations of trying to apply any particular rule of thumb about who should pay because there are problems in trying to rigorously implement either of the rules. One problem involves lack of information. For diffuse environmental problems, we often do not know in detail which farmers are the main sources of the problem. Often, the cause-and-effect relationships between any environmental works and environmental outcomes are unknown or highly uncertain. For environmental issues that have nonmarket benefits and costs (e.g., existence values), we do not know who the beneficiaries are. Lack of information means that if a government sets out to implement either of the rules as being the most fair, it could easily end up with a distribution of benefits and costs that deviates a long way from the chosen rule. Market equilibration also means that the distribution of benefits and costs may diverge from that originally intended by government.

Budgetary Implications

The choice of the beneficiary-pays or polluter-pays mechanism has direct implications for government expenditures. Payments require governments to pay out, and the degree of political support for the subsidization of BMP adoption policies will decline with fiscal austerity pressures. The pressures can result in a reduced-scale program that reduces either the extent of the cost share or the length of time for which funds are available and the subsequent effectiveness of the incentive. For example, payment schemes for a perennial land management program that provide for an initial phase of payments (e.g., three to five years) but not for ongoing payments will likely result in a conversion back to annual crop production unless the actions being promoted have positive private benefits and the payments are being used to encourage farmers to learn about them. In other circumstances, incentive-based schemes need to allow for ongoing costs in the long run if the benefits are to be sustained.

Direct controls are probably less costly to government than payment incentive schemes, although they do incur transaction costs as discussed in the following subsection. Other polluter-pays mechanisms can even generate public revenue, which increases the attractiveness of such options for governments facing budgetary pressures. Taxes on emissions or polluting inputs are a direct source of revenue, while permit schemes can also bring in money provided the initial tradable rights are auctioned rather than allocated through a grandfathering clause. While such mechanisms place less stress on government budgets than incentive-payment schemes, Goulder and Parry (2008) note that other fiscal interaction effects can result in significantly higher public costs depending on the price distortion caused (tax-interaction effect) and on whether the revenue-recycling effect is exploited.

Transaction Costs

The costs associated with the design and implementation of a policy instrument can alter the relative net benefits of alternative mechanisms and ultimately the appropriate policy choice. McCann et al. (2005) define transaction costs for the creation and use of an environmental policy to include: (1) research, information collection to define the problem; (2) enactment of legislation; (3) policy design; (4) implementation cost; (5) support and administration of policy; (6) compliance monitoring; and (7) enforcement. The timing of costs over the lifetime of the policy and the incidence of costs for different groups (e.g., public or private parties) vary with the category and type of policy mechanism. In addition, the costs of creating and using a policy depend on the nature of the transaction for the externality, the characteristics of the farmers affected and of the policy administrators, and the current institutional arrangements in place (Coggan et al., 2010).

Transaction costs tend to be lower for design-based than performance-based mechanisms for most agricultural externalities. Since many of the agricultural emissions discussed earlier tend to travel by diffuse and indirect pathways potentially over long periods of time to the affected resource, it is costly to determine cause and effect, thereby raising the transaction costs of performance-based mechanisms. For example, the failure of many water quality–trading systems in agricultural watersheds is due in part to the high transaction costs associated with predicting emission loads from each farmer (Shortle, 2013; Woodward, 2000). While targets tend to be clearer and more easily understood for design-based instruments, transaction costs with these policies increase if the externality is significantly influenced by factors such as the timing or method of application that are difficult to observe.

For either the performance or design-based mechanisms, transactions costs also depend on other physical factors related to the generation of the externality, aside from the ability to measure and observe the environmental target. The extent of the change in management practices required influences not only abatement costs (private net benefits in Figure 1), but also the transaction costs of either instrument. In situations where the required alteration in current practice is relatively large, direct controls are more likely than an incentive scheme to result in strong resistance and the need for expensive monitoring and enforcement. The greater the number of farmers involved, the greater the costs of implementing the policy and potentially the greater the lobbying efforts to prevent its implementation (McCann, 2013). Regardless of whether a direct control or payment scheme is being considered, transaction costs will increase with the scale of the problem as more individuals and more legislative jurisdictions are involved to resolve the externality, making it more difficult to coordinate efforts.

The relative transaction costs of payment schemes versus direct controls are also influenced by current institutional arrangements. Previous policy decisions can either enable or constrain the design of either mechanism. The promotion of practices to improve farmer profitability has been an integral part of extension programs globally, and expanding these extension efforts to include the demonstration and education of practices to reduce externalities can be incorporated with relatively low transaction costs. Similarly, payments for the adoption of easily observable practices can be readily incorporated into existing agricultural support programs. The existence of programs such as crop and net revenue insurance also influences the political feasibility of the mechanism choice; the public in some countries continues to be in favor of support for family farms, making it more likely that the environmental policy instrument will similarly involve a payment scheme. Aside from their political acceptability, taxing polluting inputs or establishing tradable permit schemes may require the establishment of new administrative departments to implement and enforce.

Distributional Implications

In addition to whether the beneficiary or polluter pays, there are several distributional considerations for each incentive mechanisms. The benefits and costs can differ across farm types, sectors, regions, members of society, and generations.

The physical and management characteristics vary across farm operations, a heterogeneity that has implications for the effectiveness of uniform control measures and for who bears the costs of addressing the externality. Damages from agricultural production within a given location can vary with the proximity to a watercourse or soil characteristics such as soil type and topography. The private net benefits of BMPs in reducing damages will also vary across those same farms owing to management ability and operation size, and this heterogeneity is unlikely to be correlated with the environmental performance of the farms. The variation implies that there can be a difference in the choice of BMP or differences in the most appropriate policy mechanism for alternative farms in the Public:Private Benefits Framework (Figure 1). The greater the heterogeneity, the greater the efficiency associated with a targeted approach—but also the greater the transaction costs.

The choice of mechanism in other jurisdictions can influence the distributional implications of alternative policies and the subsequent choice by a given municipality. Interest in cap-and-trade markets for carbon among states in the United States and provinces in Canada has increased with the implementation of a permit market in California. However, governments have been reluctant to impose a polluter-pays mechanism on export-oriented sectors, fearing it will negatively impact the competitive advantage of that sector. The fear of firms moving to areas with the laxest environmental regulations, and thus lower production costs, is embodied in the pollution haven hypothesis (Copeland, 2008). There is some evidence of spatial movement in the U.S. livestock industry owing to differences in regulatory stringency, but the major reason for the changes appears to be agglomeration economies (Herath et al., 2005). There is also growing evidence of support for the Porter Hypothesis that stricter environmental regulation does not negatively affect a sector or region but rather leads to more innovation and potentially stronger business performance (Ambec et al., 2013).

The market will also have an influence on the distribution of benefits and costs, irrespective of government wishes. For example, if farmers’ production costs go up as a result of legal requirements to protect biodiversity, the farmers may or may not be able to pass on the increase to consumers of their products, a group that may include many beneficiaries of the new law. If the price elasticity of demand is high, farmers lose more than they gain by attempting to pass on the extra costs. In an unregulated market, the distribution of costs between farmers and consumers is completely outside government control as it depends on the price elasticities of supply and demand, and these depend on producers’ cost structures and consumers’ preferences, not on government policy. The capacity to pass on costs also depends on the market structure in intervening steps of the supply chain. Costs could potentially be absorbed within the chain such that neither polluters nor beneficiaries pay.

Multiple Policy Goals

The public assessment of what is fair and equitable has justified the use of government funds for farm support programs and for payment schemes to incentivize the alteration of current management practice. The joint objectives of enhancing both farm income levels and environmental performance may not necessarily coincide.

Agricultural support programs may have contributed to the externality issues associated with modern farming by enhancing the average and reducing the variability of returns for more erosive and chemically intensive row crops such as corn and soybeans, thereby encouraging intensification and specialization of farms (Tilman et al., 2002). Claassen et al. (2017) found that crop insurance influenced crop choice but had a small impact on agricultural pollution, while other studies on crop insurance suggest that it depends on whether the input associated with the residual (i.e., fertilizer, pesticide) is a risk-increasing or risk-decreasing input (Smith & Goodwin, 1996).

When an agri-environmental program is combined with an insurance program, the intended impacts from one policy may be offset by the other. For example, Goodwin and Smith (2003) estimated that the Conservation Reserve Program (CRP) reduced soil erosion significantly but half of the reduction was offset by increased erosion caused by other farm support programs, not including crop insurance. Similarly, Miao et al. (2016) suggest that the cost effectiveness of CRP could be improved if the interactions with crop insurance programs were considered.

Practices meant to reduce the targeted environmental externality of a given policy could also positively or negatively address another externality. Land management efforts, such as cover crops, that improve soil health tend to reduce several pollutants such as soil sediment and nutrient runoff. The result is that often payment schemes are used to support the adoption of practices considered to have multiple benefits for environmental health. However, the cost effectiveness depends on ranking and prioritizing the externality. DeLaporte et al. (2010) found that the cost and location of wetlands to be preserved depended on whether the ecological goal was to improve water quality through sediment abatement or to provide wildlife habitat. The policies may not result in synergistic improvements in environmental health but instead lead to conflicts. For example, policies to support practices for reducing nitrates in groundwater could result in higher ammonia levels to the atmosphere, or pesticide reduction policies could lead to higher soil erosion levels from the need for more tillage (Finger et al., 2017).

The Role of Uncertainty

Uncertainty influences the choice of incentive mechanism, and this choice in turn has an impact on the type of uncertainty that arises (Goulder & Parry, 2008). The relationship between the actions of an individual farmer and the subsequent externality is not known with certainty and is subject to stochastic events, like rainstorms, that are outside of the farmer’s control. This increases the transaction costs of designing an instrument to account for all potential contingencies (McCann, 2013) and decreases the political feasibility of imposing a regulation on a farmer, whose actions are not completely responsible for the damages.

Uncertainty also influences a producer’s response to a directive. For example, farmers in many places tend to apply more than the recommend rate of fertilizer due to them, perceiving that the benefits of overapplication in good years outweigh the higher-than-average cost in other years (Rajsic et al., 2009). The necessary subsidization to encourage a change in behavior is subsequently greater than suggested by a simple partial budget of abatement costs with no uncertainty. Shortle (2013) notes that reducing uncertainty about the predictions on decisions made by individuals and the ecological outcome of those choices is important for improving the design of incentive mechanisms by policymakers.

The nature of uncertainty varies with the type of agri-environmental policy. Actual prices under a tradable permit scheme are initially unknown, although price floors are often established (Goulder & Parry, 2008). In contrast, the price incentives under a subsidization scheme are known but the extent to which behavior is modified and subsequently its impact on environmental health is unknown. The level of adoption of the desired BMPs under a direct-control mechanism will be known reasonably well beforehand, although its actual influence on the externality is subject to the random factors discussed previously. The relative efficiency of a price or quantity approach depends on the relative elasticities of marginal abatement and marginal damages (Weitzman, 1974).

Summary

The type and scope of externalities associated with agriculture vary significantly by location, but the majority of negative externalities tend to involve numerous diffuse and heterogeneous contributors. The impact of their actions on environmental quality is indirect and stochastic, making it difficult to assign liability. Subsequently, agri-environmental policies do not tend to target the externality directly but instead focus on observable practices. These design-based policies are not as efficient but have lower transaction costs than performance-based instruments.

The design-based policies to incentivize farmers to alter their management practices tend either to provide payments to encourage adoption or to impose direct regulations requiring farmers to undertake certain practices. These two incentive-based mechanisms are recommended where the private incentives conflict with the public interest and only where the private incentives are not so strong as to outweigh the public benefits. Both the provision of payments for BMPs and a regulatory requirement for use of a BMP alter the incentives faced by farmers, but they do so in different ways, with different implications and consequences. The biggest differences between them probably relate to equity/distributional outcomes and politics rather than efficiency. Governments often seem to prefer to employ beneficiary-pays mechanisms in cases where they seek to alter farmers’ existing practices and polluter-pays mechanisms when they seek to prevent farmers from changing their current practices to something worse for the environment. In other words, the status quo commonly defines de facto property rights.

Past technological developments have provided practices, such as conservation tillage, that reduce the environmental externality but are also profitable for the farmer. The digitalization of agriculture is the next technological revolution facing the sector and could provide significant opportunities to improve the environmental performance of farm operations. The digital revolution has the potential to help farmers produce more food on less land and with fewer inputs. Whereas previous technological advances in agriculture increased productivity by creating uniform management systems, this new digital revolution uses the tools of Big Data analytics to give farmers the ability to tailor management options to the specific requirements of numerous zones, which can be as small as an individual plant. In addition to reducing input levels and identifying unprofitable management zones to set aside, the technology could also alter the transaction costs of the policy options. Targeted approaches may become cost effective particularly since many of the environmental issues are associated with only a few producers (Leslie et al., 2017). Costs of measurement can be reduced with digitalization, thereby altering the cost effectiveness of the instruments and potentially allowing for the use of performance-based incentive mechanisms.

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