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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution 8 Water Quality Goals KEY POINTS IN CHAPTER 8 This chapter discusses how the setting of water quality goals can be used for combating nutrient over-enrichment problems in coastal areas and finds: To design effective approaches to mitigating nutrient over-enrichment, decisionmakers must understand the physical and ecological processes at work, outline clear management goals, set specific targets, and develop a range of possible policy approaches or management tools that are suitable to the site and its problems. To make efficient use of available resources, managers should adopt policies that ensure that targets will be met at the lowest possible cost. In many cases, control costs will vary across sources and, if equally effective, the total cost of meeting the target will be lowest if the lowest cost sources are controlled first. In designing policies to achieve nutrient reductions, decision-makers will need to choose between voluntary approaches and mandatory controls or financial penalties. Each approach has advantages and disadvantages, and managers must assess how successful a given approach is likely to be in their specific context. Voluntary changes in behavior can be difficult to motivate. Providing information and education is not always effective, but it is relatively inexpensive and non-controversial. Providing subsidies designed to reduce nutrient inputs can be effective, but requires funds that are generally raised via taxes, which may impose a cost on society. Also, subsidies can inadvertently encourage pollution, because polluters are not required to pay the full costs of their activities, which in turn can lead to lower product costs, then higher product demand, and ultimately increased pollution to meet that demand.
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution Some of the shortcomings of both regulatory and tax-based approaches can be overcome with the use of marketable permits. A careful examination of the effectiveness of this approach where it has already been implemented should be undertaken. In many instances, managers may find that a well-formulated mix of incentives (voluntary approaches) and disincentives (mandatory or punitive approaches) works better than either approach alone. Managers might increase the likely success of a voluntary approach by making it clear that, if the voluntary approach is not successful, an approach based on disincentives will be adopted. An information clearinghouse should be established that provides local managers with information about the cost and effectiveness of alternative source control methods, the effectiveness of alternative policy options, and the policy experiences that other managers have had in attempting to control nutrient over-enrichment. This information should emphasize the role of site characteristics in determining effectiveness and costs. In developing an effective strategy for mitigating the effects of nutrient over-enrichment one must understand the physical and ecological relationships that determine the extent and causes of nutrient over-enrichment, along with societal objectives and behavioral responses. Societal objectives determine goals that a management strategy will strive to achieve and the benchmark against which it will be evaluated. Behavioral responses ascertain how the various parties contributing to nutrient over-enrichment are likely to respond to different policies designed to affect that behavior. Managers must anticipate and understand these responses as they choose among policy alternatives, since the responses will determine the effectiveness of any given policy. This chapter discusses issues that arise both in setting goals for nutrient over-enrichment management strategies and in choosing among policy alternatives. The appropriate set of policies for any given estuary will depend on the nutrient sources for that estuary. For example, if the main nutrient source is agriculture, a set of policies designed to promote the adoption of best management practices is required. These can be implemented at the local, regional, or national level. Alternatively, if atmospheric deposition is the main source of nutrients, policies that reduce atmospheric emissions of nitrogen are needed. Since the source of atmospheric nitrogen is often outside the local jurisdiction governing the estuary, policies to combat this nutrient source must be implemented at the regional or national level. Because both the susceptibility and the specific policy needs of any given waterbody are site-specific, this chapter does not attempt to prescribe specific water quality goals or policy choices for adoption by local managers. Rather, it is intended to provide managers with an improved understanding of the factors that should be considered in setting goals and making policy choices. With this understanding, managers will then be able to begin the process of crafting a
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution management strategy that addresses the needs and challenges of waterbodies in their jurisdiction. SETTING GOALS Choosing Targets The most common approach to setting water quality goals to combat nutrient-related problems is to select target levels of ambient concentrations in the receiving body, nutrient loadings to the receiving body, or resource stocks (e.g., acres of submerged aquatic vegetation [SAV] or marine populations). Concentration levels are not always useful indicators of the eutrophic state of an estuary, since they reflect both nutrient inputs to the estuary and the response of the estuary. For instance, nutrient concentrations can be low in a highly eutrophic estuary if the nutrients are rapidly taken up by phytoplankton. Conversely, nutrient concentrations can be high in a non-eutrophic estuary if factors such as short residence times or low light availability (from deep mixing and high turbidity) make the estuary non-susceptible to eutrophication. For this reason, targets based on primary productivity (Nixon 1995), chlorophyll, or phytoplankton biomass (NRC 1993a) are likely to be better indicators of an estuary’s eutrophic state and hence better measures of whether overall water quality goals (i.e., desired reductions in eutrophication) are being achieved. Setting policy goals involves not only choosing a target indicator but also setting a target level for that indicator. There are a number of different bases that can be used to set target levels. One possibility is to seek to increase/decrease the indicator value by some arbitrary amount (e.g., 25 percent) or to restore it to its level during some previous period. For example, for Chesapeake Bay, an arbitrary goal of reducing controllable sources by 40 percent was set. Similarly, managers might seek to restore SAV acreage to some historical level. Alternatively, the target level might be determined by a specific use that is desired (e.g., a “fishable/swimmable” criterion). Setting target levels in either of these two ways typically focuses on the benefits of improvements, or, similarly, on what level of improvement is deemed “feasible.” In such cases, the target level is chosen without an explicit regard to the cost of achieving the improvement, although that cost is implicit in the definition of “feasibility.” Similarly, the benefits gained by meeting the target based on “feasibility” are not always clearly articulated. Another way is to consider explicitly both the costs and benefits of alternative targets. For example, a target level of SAV acreage could be based not on an historical level but on a comparison of the cost and
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution benefit of achieving the restoration. Such a comparison requires a measure of the benefit of restoration or a measure of the cost of degradation. Unfortunately, these costs can be difficult to estimate. Although it may be relatively easy to calculate the commercial value of a resource that is harvested and sold in the market (e.g., a commercial fish stock), the non-commercial (or non-market) value of natural resources is inherently difficult to estimate. Nonetheless, techniques exist for valuing non-market goods such as wetlands, water quality, and wildlife populations, and these techniques have been applied to the valuation of estuaries (Chapter 4). With estimates of both the benefits and costs of achieving improvements in water quality or resource stocks, an economically efficient target level for improvement can be identified. Improvements would be sought up to the point where the cost of any additional improvement would exceed the benefit of the additional improvement. However, since the ability to estimate benefits of improving winter quality remains imperfect and imprecise, the use of this approach alone can be difficult. Establishing Criteria and Standards Targets usually fall into two categories: water quality criteria and water quality standards. McCutcheon et al. (1993) differentiate the two as follows: A water quality criterion is that concentration, quality, or intensive measure (e.g., temperature) that, if achieved or maintained, will allow or make possible a specific water use. [For the toxic substance,] a criterion may be a concentration that, if not exceeded, will protect an organism, aquatic community, or designated use with an adequate degree of safety.1 A criterion may also be a narrative statement concerning some desirable condition. While water quality criteria are often the starting point in deriving standards, criteria do not have a direct regulatory impact because they relate to the effect of pollution rather than its causes. A water quality standard is the translation of a water quality criterion into a legally enforceable ambient concentration, mass discharge or effluent limitation expressed as a definite rule, measure, or limit for a particular water quality parameter. A standard may or may not be based on a criterion. Standards may differ from criteria for a variety of reasons, including natural impairment of water quality even in the absence of anthropogenic 1 As noted above, nutrient concentrations do not alone provide adequate criteria for control of eutrophication in estuaries. However, the oxygen concentration of an estuary could be used, setting a concentration below which oxygen should not fall. Other appropriate criteria might include chlorophyll concentrations or nutrient inputs.
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution pollution, the perceived importance of a particular ecosystem, or the degree of safety required for a particular waterbody (McCutcheon et al. 1993). While criteria are typically defined relative to ambient water quality, standards may take different forms. Ambient standards often are based on the establishment of threshold values for a particular contaminant, and they consider the intended use of the waterbody, as well as its ability to assimilate wastes. Effluent standards limit the amount of material that may be discharged regardless of the size of the receiving waterbody or the intended use of its waters. Effluent standards are often technology-based and may be imposed even if the level of contamination is less than that required to achieve ambient water quality standards (McCutcheon et al. 1993). When a receiving waterbody is affected by a discharge, the standard, be it ambient or effluent, must govern the discharge or loading of wastes into the water. Even for large rivers, loadings over the entire watershed may impact the estuarine or coastal water quality. Traditionally, water quality standards have been absolute numbers in the sense of a concentration or discharge of a toxic substance that may not be exceeded, or an oxygen concentration that must be maintained. However, water quality in a given waterbody can fluctuate as a result of random factors, such as weather and uncertainties in hydrologic processes. When natural variation is important, it is still possible to implement probabilistic standards—standards that, for example, specify that a given concentration of a toxic substance is not to be exceeded more than once, on the average, during a certain number of years (EPA 1991b).2 Probabilistic standards account for extreme hydrologic events that rarely create excessive nonpoint source loads, and are more reasonable when receiving-water impacts are driven in large part by these factors. Other impacts of nutrient enrichment could also lead to a standard, if criteria can be established through scientific studies and modeling. Eutrophication impacts were discussed in Chapter 4. They include increased primary productivity; increased phytoplankton biomass; reduction in water clarity; increased incidence of low oxygen events (hypoxia and anoxia); changes in the trophic structure, trophic interactions, and trophodynamics of phytoplankton, zooplankton, and benthic communities; damage to coral reefs; fish kills; reduced fisheries production; and decreased biotic diversity. In general, the public must accept the need for improved receiving-water quality and prevention measures. If the public does not perceive a problem, it is unlikely that elected officials will pur- 2 EPA’s total maximum daily load (TMDL) process allows the use of probabilistic standards (EPA 1991b). See further discussion below.
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution sue this issue or that agency staff will have the resources or authority to implement solutions. Finally, as with any other complex water quality problem, many other factors must also be considered in setting nutrient-based water quality standards, including: Are damages subject to threshold effects or are they continuous? In some cases, impacts are gradual up to some critical loading, at which point the receiving water may react in a way that greatly increases algal growth. In other cases, impacts remain gradual throughout. (Susceptibility is discussed in Chapter 5.) During which seasons are impacts the greatest? Warmer water during the summer often leads to worsening of dissolved oxygen levels because of lower saturation concentrations and higher rate constants, combined with greater stratification and more light. If there is a low-flow season, dilution of wastes entering an estuary may be lower. Year-to-year variability in climate is also a driving force. Critical conditions can usually be determined by simulation over periods of at least a year. Are sudden discharges (e.g., due to a thunderstorm) controlled? In some settings, a heavy rainfall during dry weather may lead to intense, temporary nonpoint source loading. What are the flushing and mixing conditions in the estuary or coastal water? Exchange with the open ocean or sometimes with wetlands may mitigate the impact of heavier loadings. Effects will differ for every coastal water. What time scales are involved? Control efforts are likely to take a significant amount of time to become effective in reducing an existing negative impact. It is likely that improvements due to reduced nutrient loadings will be felt only over a period of several years because of nutrient storage in the system, especially in sediments, although in some cases some systems may show a rapid response. Current Criteria and Standards Most water quality criteria in the United States are based on the “Gold Book” (EPA 1987), in which criteria are given for over 100 constituents, most of which are heavy metals and organic chemicals. Criteria may differ for fresh and marine waters and may be based on toxicity to various aquatic organisms or hazard to human health (e.g., drinking water standards). Another current approach to regulating water quality, defined in section 303(d) of the Clean Water Act, is an approach known as total
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution maximum daily load (TMDL). This method, which was refined in later regulations, is a water quality-based standard. It strives to assure water quality through a series of steps that, in effect, require a watershed scale approach (EPA 1999a). For problems of eutrophication in estuaries, this emphasis on water-quality-based standards was recommended by a previous National Research Council (NRC) report (NRC 1993a) and is endorsed by this committee. Under proposed 1999 modifications to the TMDL process, a TMDL must contain the following minimum elements (EPA 1999a): the name and geographic location of the impaired or threatened waterbody for which the TMDL is being established; identification of the pollutant for which the TMDL is being established and quantification of the pollutant load that may be present in the waterbody and still ensure attainment and maintenance of water quality standards; identification of the amount or degree by which the current pollutant load in the waterbody deviates from the pollutant load needed to attain or maintain water quality standards; identification of the source categories, source subcategories, or individual sources of the pollutant for which the wasteload allocations and load allocations are being established; wasteload allocations for the pollutant to each industrial and municipal point source; for discharges subject to a general permit, such as storm water, combined sewer overflows, abandoned mines, or combined animal feeding operations; pollutant loads that do not need to be reduced to attain or maintain water quality standards; and supporting technical analyses demonstrating that wasteload allocations, when implemented, will attain and maintain water quality standards; load allocations, ranging from reasonably accurate estimates to gross allotments, to nonpoint sources of a pollutant, including atmospheric deposition or natural background sources; and supporting technical analyses demonstrating that load allocations, when implemented, will attain and maintain water quality standards; a margin of safety expressed as unallocated assimilative capacity or conservative analytical assumptions used in establishing the TMDL (e.g., derivation of numeric loads, modeling assumptions, or effectiveness of proposed management actions that ensure attainment and maintenance of water quality standards for the allocated pollutant); consideration of seasonal variations and environmental factors that
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution affect the relationship between pollutant loadings and water quality impacts; an allowance for future growth, if any, which accounts for reasonably foreseeable increases in pollutant loads; and an implementation plan, which may be developed for one or a group of TMDLs. In most cases, item 4 (identification of source categories, etc.) of this process virtually mandates a watershed approach since waters are impaired by multiple dischargers and pollutants, and these derive, to a considerable extent, from nonpoint sources distributed over broad regions. The Environmental Protection Agency (EPA) recommends that TMDLs be developed on a regional basis, that is, by watershed. A virtue of the TMDL approach is that it is flexible and considers water quality to be a function of an extensive range of sources distributed across the landscape. It is so flexible that physical and biological stressors, like water temperature and habitat alteration, can be considered within the same management framework (NRC 1999a). The TMDL approach is promising for control of pollution, including nutrients, but it can be hampered by data gaps. For instance, development of a TMDL presupposes that a waterbody has been classified as water quality impaired, that its condition has been ranked and prioritized with respect to other impaired waters within a state, and that standards for specific contaminants have been established. Development of TMDLs requires understanding of point and nonpoint sources; the processes that influence the magnitude, timing, transport, and attenuation en route of pollutants; and how those pollutants affect aquatic biota. The TMDL procedure is fairly site-specific. Thus, the management approach tends to be tailored to the needs of a specific watershed or receiving body. Because of data gaps and limitations in knowledge of the structure and function of watershed ecosystems, and because the approach is still relatively new, development of TMDLs has proceeded slowly. Table 8-1 summarizes current nutrient criteria for coastal states by EPA region (EPA 1998c). As discussed in Chapter 7, information on the ambient concentration of nutrients in watersheds and rivers is important for calculating the load that downstream coastal water bodies will receive from these sources. However, resistance to concentration-based standards by some coastal states is understandable, given their limited utility as a measure of waterbody impairment. Ambient concentration of nutrients in receiving waters rarely reflects the degree to which the body has been impacted by nutrient over-enrichment. For example, if an estuary is nitrogen limited, primary productivity will be stimulated by nitrogen loading. This increased primary productivity will then remove nitrogen
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution TABLE 8-1 Region/State Nitrate Ammonia Total Nitrogen Total Phosphorus Region 1 Connecticut 3 Maine 7,3,8 Massachusetts 2 2 2 New Hampshire 2 Rhode Island 3 3 3 Region 2 New Jersey 2 2 9,3 New York 2 Puerto Rico 9 9 8,2 Virgin Islands 8,9 Region 3 Delaware 2 2 2 District of Columbia Maryland Pennsylvania Virginia 4 Region 4 Alabama Florida 2 7,2 Georgia 3 Mississippi North Carolina 3 South Carolina 2 2 3 Region 6 Louisiana 2 2 2 Texas 2 2 2 Region 9 American Samoa 2 2 1,9 1,9 California 1,5 5 1,2 1,6,7 Guam 2,7 5 2 2,7 Hawaii 1,9 1,9 1,9 1,9 Nevada 5 2,5 1,7,9 1,7,9 Northern Mariana Islands 7 5 7 7 Trust Territories of the Pacific Islands 2 2 7 7,9 Region 10 Alaska Oregon Washington TABLE 8-1 Summary of existing water quality criteria and standards for nutrient over-enrichment for coastal states and territories (EPA 1998c). As of 1999, 21 states and territories had proposed water quality criteria for nutrients, with significant differences in the nutrients addressed and whether the criteria are narra
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution tive or quantitative. Blank entries indicate that no criterion for the nutrient has been specified by the state. Key: (1) Site-specific numeric values for ambient nutrient levels; (2) narrative criteria related to natural conditions, eutrophication and nutrient over-enrichment for nitrate, ammonia, inorganic nitrogen, total nitrogen, or total phosphorus; (3) narrative criterion that is not related to natural conditions, eutrophication, or nutrient over-enrichment issues; (4) numeric values for effluent nutrient levels; (5) numeric values related to public health (nitrate) or aquatic toxicity (ammonia); (6) habitat-based numeric values for ambient nutrient levels; (7) water use classification- or water use designation-based numeric values for ambient nutrient levels; (8) state-wide numeric values for ambient nutrient levels; and (9) waterbody-based (streams, rivers, lakes, estuaries, coastal/oceanic waters) numeric values for ambient nutrient levels. from the water column at a high rate and tie it up in organic matter. Thus, the ambient nitrogen concentration in the water column may never rise significantly, or remain elevated long enough to be observed, even as eutrophication of the body takes place (Box 8-1). The TMDL approach avoids this limitation by directly addressing nutrient loading. As the TMDL approach evolves in support of controlling nutrient problems in coastal waters, it will need to recognize the variation among estuaries and follow a consistent classification scheme (Chapter 6). Criteria endpoints are likely to vary by type of estuary (e.g., some might address seagrass extent, others chlorophyll, and others dissolved oxygen). EPA recently developed nutrient standards on a regional or watershed basis (EPA 1998e). The major elements of this strategy include: use of a regional and waterbody-type approach for the development of nutrient water quality criteria; development of waterbody-type technical guidance documents (i.e., documents for streams and rivers; lakes and reservoirs; estuaries and coastal waters; and wetlands) that will serve as user manuals for assessing trophic state and developing region-specific nutrient criteria to control over-enrichment; establishment of an EPA national nutrient team with regional nutrient coordinators to develop regional databases and to promote state and tribal involvement; EPA development of nutrient water quality criteria guidance in the form of numerical regional target ranges that EPA expects states and tribes to use in implementing state management programs to reduce over-enrichment in surface waters (i.e., through
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution the development of water quality criteria, standards, EPA’s National Pollutant Discharge Elimination System permit limits, and TMDLs); and monitoring and evaluation of the effectiveness of nutrient management programs as they are implemented. The major focus of this strategy is the development of waterbody-type technical guidance and region-specific nutrient criteria by the year 2000. Once the guidance and criteria are established, EPA will assist states and tribes in applying numerical nutrient criteria to water quality standards by the end of 2003. CHOOSING A POLICY APPROACH Once water quality or other resource-based targets have been set, managers must decide which policy approaches to use for achieving those targets and the details regarding implementation. Some approaches are based on voluntary action, while others involve mandatory controls or the use of financial or other penalties (e.g., taxes) to induce desired behavioral changes. In this section, the strengths and weaknesses of alternative approaches are discussed. Evaluation Criteria In choosing among alternative approaches to reducing eutrophication or other effects of nutrient over-enrichment, it is important to specify the criteria used to rank the alternatives. The type of criteria typically used (e.g., Bohm and Russell 1985; NRC 1993a) are: cost-effectiveness; dynamic adjustment (flexibility, adaptability, and innovation incentives); and distributional impacts and fairness. Cost Effectiveness Given a predetermined target or goal for a given response (e.g., reduced eutrophication), cost effectiveness is achieved when that target is met at the lowest possible cost (Bohm and Russell 1985; Baumol and Oates 1988). Applying this principle requires both a definition of effectiveness and a mechanism for ensuring that, other things being equal, low cost sources are controlled first. In defining effectiveness, a number of issues arise. First, the criterion must be based on measures of water
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution ity of reducing loadings, for example, through improved energy efficiency, and such improvements are cost-effective; (2) produce products where demand is sensitive to their environmental characteristics or energy efficiency; (3) perceive public relations benefits from participation in voluntary programs; or (4) fear the imposition of mandatory controls if voluntary approaches are not successful in achieving desired reductions in loadings or fossil fuel use. Mandatory Policies While information-based policies and subsidies rely on voluntary changes in polluting behavior with no long-term net cost to polluters, mandatory policies dictate behavioral changes or payments based on polluter choices. Irrespective of whether the mandatory policy takes the form of command-and-control regulation, taxes, or fees, it puts the burden and the associated cost of pollution control on the polluters (Box 8-3). Command-and-Control Regulations Command-and-control regulations can take a number of forms, including mandatory limits on emissions of a pollutant (e.g., NOx emission caps or nitrogen effluent limits), required investment in pollution control equipment (e.g., use of best available control technologies), or required use of specified production practices (e.g., reduced tillage). To be cost effective, regulations must be designed to ensure that pollution reductions are achieved in the least costly way. Historically, regulations have not always been designed with this goal in mind, and they have been criticized for their high costs (Hahn 1994). Environmental regulations have relied heavily on the use of technology standards, which require installation of a particular type of pollution control equipment and are generally not cost effective. This standardized, “one-size-fits-all” approach deters firms from developing and taking advantage of alternative, less costly technologies and methods of reducing emissions. More recently, the nation has moved toward greater reliance on performance standards, which grant polluters the flexibility to meet standards in a variety of ways, and this is expected to lead to greater cost effectiveness (Besanko 1987; Burtraw 1996) and encourage innovation. With technology standards, firms have no incentive to develop less costly approaches to pollution reduction, since the regulation does not allow them to benefit from such improvements. With a performance standard, any reduction in the cost of meeting the standard (through, for example, an innovation in pollution control techniques) generates direct benefits
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution BOX 8-3 Maryland Tries Mandatory Nutrient Management In 1998 Maryland passed its Water Quality Improvement Act (WQIA), perhaps the most comprehensive farm nutrient control legislation in the country. The law marks a transition from voluntary to mandatory nutrient management, and it brings new attention to phosphorus as a nutrient of concern. Under WQIA, by 2005 all agricultural operations with annual incomes greater than $2,500 or more than eight animal units (one animal unit equals 1,000 pounds of live weight) must implement nutrient management plans that consider both nitrogen and phosphorus application rates. In the past, when animal manure or sludge was applied, the amount of recommended materials was based on crop nitrogen needs. However, because the amount of phosphorus in manure is generally high relative to nitrogen and the nutrient needs of growing crops, this practice resulted in substantial excess application of phosphorus. Although it was long thought that controlling erosion controlled phosphorus loss, research has shown that, even without erosion, runoff from soils with excessive phosphorus levels can contain high levels of dissolved phosphorus. The law allows at least three approaches to phosphorus control. Farmers can test their soil and follow recommendations to match agronomic and environmental needs, although this approach might greatly restrict phosphorus application on soils with optimum to slightly excessive levels without considering other site-specific factors that affect phosphorus loss. Farmers also can establish “critical” soil test values that limit phosphorus application, meaning that a level could be established at which only as much phosphorus as the crop removes could be applied, while for soils at some higher level no additional phosphorus could be applied. Scientists have objected to both approaches, since their research indicates that many site-specific factors influence the potential for phosphorus loss. Instead, they have proposed the use of a phosphorus site index. This phosphorus site index is a generalized national index that has been developed and is now being adapted by the University of Maryland for possible use in Maryland. The index evaluates slope, runoff potential, proximity to surface water, soil phosphorus levels, and fertilizer and manure application rates and methods; it thus allows site-specific assessments and comprehensive evaluation of potential environmental impacts without restricting phosphorus application to low-risk sites. To help farmers meet WQIA requirements, Maryland has committed $800,000 per year for at least three years for agricultural research and education programs, which could include research and extension programs on composting, analysis of the pilot transport program, animal nutrition management, development of a phosphorus index, and phosphorus dynamics in soils (EPA 1999b). for the firm in the form of reduced compliance costs. Thus, firms have an incentive to innovate and adopt new, more efficient techniques. Although greater reliance on performance standards rather than technology standards should lead to lower costs for individual polluters, achieving an aggregate target level of water quality in a watershed
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution involving multiple polluters at least cost is more complex. Each source must meet its required reduction at least cost and have the required reductions allocated efficiently across sources. The total cost of meeting an aggregate abatement target is minimized when the required reductions are allocated so that each source faces the same incremental cost from additional abatement (NRC 1993a). Unfortunately, if abatement costs differ across sources, this implies different required reductions for different sources. Such differential regulation can be both administratively complex and politically difficult to implement. Who pays to compensate for environmental damages is another difficult issue. Under regulations, polluters pay only for the cost of complying with regulations and not for the damages that any remaining pollution causes. As a result, the price of their products does not reflect all the associated costs of production, including both market and non-market costs. With the product price artificially low, consumption of those products tends to be high. For example, if agricultural producers comply with regulations, but nutrient runoff still occurs, that runoff could still generate costs for society (e.g., increased eutrophication) that the farmer did not consider when making production and pricing decisions. If these costs were reflected in the product prices, prices would rise and the demand for those products would adjust to reflect the full cost of production. Thus, those who bear the environmental costs of the production would no longer be implicitly subsidizing consumers of agricultural products. However, higher agricultural prices could cause hardship both to marginal farmers who might be forced out of business and to low-income consumers. The use of regulation is consistent with the polluter pays principle, to the extent that polluters pay for compliance with the regulations. However, as noted earlier, they do not pay for any damages that result despite that compliance. There have been numerous studies of the impact of nutrient-based regulation, particularly in the context of agriculture. The type of regulation (e.g., mandated reduction in excess application or limiting animal densities) strongly affects the burden and effectiveness of regulation (McSweeny and Shortle 1989). Many studies have found that regulation is more efficient when aimed at areas or farms with the greatest pollution contributions, but this increases the administrative cost of regulations substantially, and these administrative costs may outweigh the efficiency gains from varying regulations across sources according to their pollution contributions (Mapp et al. 1994; Moxey and White 1994; Carpentier et al. 1998; Yiridoe and Weersink 1998).
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution Taxes and Fees In contrast to regulatory approaches, which mandate certain changes in behavior, taxes and fees (negative economic incentives) are designed to induce (rather than force) those changes using financial incentives. They take a variety of forms, including effluent charges, user charges, product charges, administrative charges, tax differentiation, non-compliance fees, performance bonds, and legal liability for damages (NRC 1993a). For example, the state of Florida has a coastal protection tax of two cents per barrel that is charged for pollutants (petroleum products, pesticides, chlorine, and ammonia) produced in or imported into the state. The revenue from this tax goes to the Coastal Protection Trust Fund, which is used by the Florida Department of Environmental Resources for cleaning up spills (NRC 1997b). In principle at least, taxes and fees are the negative counterpart to the positive incentives created by subsidies or cost sharing. With positive inducements, polluters receive payment for voluntarily undertaking desired behavior or investment. With negative inducements, they are forced to pay for undesirable behavior. A common feature of economic incentives is that they put a price on environmental degradation. Whether in the form of forgone subsidy or explicit tax payment, polluters pay for “consuming” (or reducing) environmental quality just as they pay for the use of other inputs, such as labor and capital. Economic incentives thus put environmental inputs on a par with other inputs used in production. As with other inputs, polluters have an incentive to use environmental inputs only up to the point where the polluter’s benefit from increased use equals the price the polluter must pay for that use. As that price rises, they face an increased incentive to reduce use of environmental inputs. One of the main advantages of pollution taxes over regulatory policies is that they are generally thought to be more cost effective. Since polluters directly benefit from any cost savings, each polluter is encouraged to reduce its emissions in the least costly way. Polluters with low abatement costs have an incentive to reduce emissions more than high-cost polluters. As a result, the allocation of emission reductions will not be uniform across sources but will be more heavily borne by low-cost sources, as required for overall cost effectiveness. However, high-cost polluters will discharge relatively more and therefore bear larger total tax burdens. In addition, polluters will have an incentive to innovate, since they will benefit directly from any resulting cost savings. Although, in principle, pollution taxes induce efficient abatement, their actual effectiveness is likely to be uncertain, at least initially. When setting a regulatory standard, authorities can be reasonably certain of the
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution resulting level of emissions (assuming polluters comply). However, when setting a tax level, regulators often cannot predict with certainty how polluters will react and what the resulting level of environmental quality will be. While the level of the instrument can be adjusted over time to ensure that targets are met, such adjustments can be costly and can generate strategic behavior by polluters (Livernois and Karp 1994). In addition, it may be costly to adjust the level of the instrument in response to changes over time in economic conditions and in the demand for improvements in environmental quality. Under tax-based instruments, polluters pay not only for the costs of any abatement undertaken but also for the remaining discharges. The resulting cost allocation is hence consistent with the polluter pays principle. Because polluters have to pay both the costs of abatement and the tax, the total cost to polluters is higher under a tax policy than it would be under a regulation leading to the same level of total discharges. While this ensures that product prices reflect the full social cost of production, the total cost may create considerable hardship both for marginal firms and low-income consumers who would be hard hit by the associated price increases. This is particularly true when, in many cases, relatively high taxes would be required to induce the desired change in behavior. The magnitude of the tax increase that would be needed to induce a reduction in discharge depends on how responsive polluters are to the tax. Numerous studies have shown that, in the case of agricultural fertilizers, farmers are not very responsive to price increases (i.e., the demand for fertilizers is generally inelastic) and hence that relatively large tax rates would be needed to ensure that environmental objectives are met (McSweeny and Shortle 1989; Heatwole et al. 1990; Johnson et al. 1991; Helfand and House 1994; Pan and Hodge 1994; Weersink et al. 1998). For example, a simulation done by Giraldez and Fox (1995) found that an ad valorem tax rate of 55 percent would have to be applied to nitrogen to induce farmers to reduce nitrogen use to satisfy drinking water standards. Even though, from an economic efficiency perspective, taxes are desirable instruments for reducing nitrogen application, the large impact on farmers may limit their appeal (Moxey and White 1994; Helfand and House 1995). Similarly, while gasoline taxes may promote reductions in fuel use and hence emissions of NOx, gasoline demand tends to be relatively unresponsive to price increases and hence large tax increases would be needed to have a significant impact on fuel consumption, particularly in the short run (Espey 1998). Such tax increases are generally viewed as politically unappealing.
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution Marketable Permits Some of the shortcomings of both regulatory and tax-based approaches to environmental protection can be overcome with the use of marketable permits (Tietenberg 1985b). A marketable permit system starts with an allocation of allowable emissions across sources, as under a regulatory approach. However, by allowing sources to trade their allocations, the final allocation (after all trades have occurred) will be such that low-cost avoiders will undertake more abatement than high-cost avoiders, and the aggregate emission reduction will have been achieved at least cost. Low-cost avoiders will have an incentive to reduce discharges below their allocation and will sell their excess permits on the market. Similarly, high-cost avoiders can buy additional permits rather than incur their high costs of pollution control. If there are a sufficient number of buyers and sellers, the resulting market for permits establishes a price for emissions that reflects the total allowable emissions (i.e., the supply of permits) and the costs of pollution abatement for all polluters (i.e., the demand for permits). Unlike other economic incentives that also establish a price, the total impact of marketable permits on environmental quality is known since the total number of permits is fixed. Thus, the use of marketable permits combines the certainty of the regulatory approach with the cost effectiveness of economic incentives. The use of marketable permits also allows a regulator to achieve any desired distribution of total costs by altering the way permits are allocated initially. Economic growth is possible without changing the total level of emissions, because new firms can simply be required to purchase permits from the market. The result is an increase in permit prices, but no increase in aggregate emissions. Numerous economic studies have shown the potential for cost savings when polluters are allowed to trade pollution permits (Tietenberg 1985b; Klaassen 1996). The success of the sulfur dioxide emissions-trading program established under the 1990 Clean Air Act amendments (for example, Joskow et al. 1998) has heightened interest in this efficient pollution control tool. While this trading program was targeted toward sulfur dioxide emissions, to the extent that it is coupled with overall reductions in fossil fuel consumption, it would also help promote reductions in NOx and the associated atmospheric deposition of nitrogen. Consideration has also been given to the use of trading programs for surface sources of water pollution, including trading between point source and nonpoint source. Such trading allows point sources to sponsor implementation of nonpoint source controls rather than further cutting back on their own emissions. Assuming nonpoint source loadings are significant and the marginal costs of nonpoint source reductions are lower than the costs of
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution additional point source pollution controls, ambient water quality goals could be met at a lower cost by substituting nonpoint source for point source reductions (Crutchfield et al. 1994). Note, however, that the trading of point and nonpoint loadings requires the establishment of an appropriate trading ratio, as well as a means of quantifying the diffuse nonpoint loadings. There have been several studies of the issue of pollution abatement trading between point and nonpoint sources. Letson (1992) has provided an economic analysis of the issue, illustrating the appeal as well as the difficulties in application of such a policy. Among the difficulties cited by Letson are monitoring, use of market power to manipulate permit price, and the distribution of the financial burden of loadings reductions. In addition, the rate at which nonpoint source abatement can be substituted for point source abatement must be established. The appropriate value of this trading ratio is uncertain because of qualitative differences between the two classes of sources. The optimal trading ratio will depend on the relative costs of enforcing point and nonpoint reductions and on the uncertainty associated with nonpoint loadings (Malik et al. 1993). Crutchfield et al. (1994) isolated several practical circumstances that facilitate source abatement trading and developed an empirical protocol to determine the extent to which they exist in coastal watersheds. Their nationwide screening analysis was not designed to locate “good” candidates for trading programs. Their goal was to rule out many coastal watersheds, thus allowing researchers and planners to better focus their water quality efforts. Several efforts are under way to implement point source and nonpoint source trading programs to improve water quality. Connecticut has applied one such program to Long Island Sound. A nitrogen-trading plan has been established to achieve reductions in nitrogen discharges cost effectively and expeditiously. The Connecticut Department of Environmental Protection anticipates that this plan will reduce the statewide bill for nitrogen removal by more that $200 million (CDEP 1998). A similar program to limit phosphorous loads exists in the Cherry Creek basin in Denver, Colorado. The Cherry Creek Trading Program involves two types of trades: authority pool and in-kind trades. In authority pool trading, phosphorous reduction credits from Cherry Creek Basin Water Quality Authority projects are allocated to a trading pool. A qualified discharger may apply to the Authority for the purchase of trade credits from the trading pool for its wastewater treatment plant. For in-kind trades, non-Authority owners of independent nonpoint source pollutant reduction facilities receive credits to be used for their own wasteload allocation or to be transferred to a wastewater treatment facility (Sandquist and Paulson 1998).
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution The Tampa Bay National Estuary Program uses a cooperative approach that resembles a watershed trading program. No actual trades take place, but some sources make pollutant load reductions that they otherwise would not have been required to make in order to offset increases occurring at other sources. This approach to watershed management may be applicable to areas where trading is technically or politically inappropriate or unnecessary (Bacon and Greening 1998). STEPS IN DEVELOPING EFFECTIVE WATER QUALITY GOALS The first step in designing an effective approach to combat the effects of nutrient over-enrichment is to understand the physical and ecological processes at work. Next, decision-makers at local or regional levels must outline clear management goals, set specific targets to achieve the overall goals, and develop a range of possible policy approaches or management tools that are suitable to the site and its problems. Targets can be based on various measures or indicators of nutrient over-enrichment or estuarine health and can take the form of general water quality criteria or more specific water quality standards. Federal development of many of the resources and research efforts called for in Chapter 2 would greatly facilitate these efforts. Once water quality goals or targets are set, managers must choose among a variety of policy approaches or management tools. To make efficient use of available resources, managers should strive to adopt policies that ensure that targets will be met at the lowest possible cost. In many cases, control costs will vary across sources and, if equally effective, the total cost of meeting the target will be lowest if the lowest cost sources are controlled first. Thus, when control costs vary, managers should not seek to achieve uniform reductions across all sources. Rather, they should target first the sources where reductions can be made at relatively low cost. In designing policies to achieve these reductions, a fundamental choice must be made between the use of a voluntary approach and the use of mandatory controls or financial penalties. There are advantages and disadvantages of each approach, and managers must assess how successful a given approach is likely to be in their specific context. In many instances, managers may find that a well-formulated mix of incentives (voluntary approaches) and disincentives (mandatory or punitive approaches) works better than either approach would work alone. Voluntary approaches that rely on moral persuasion, information, technical assistance, and possibly financial subsidies can be effective if there are sufficiently strong incentives for participation. While participation can be increased through financial incentives, local managers are not
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution likely to have the local resources to finance subsidies for participation. In addition, even if financial incentives were available, subsidies of this type are generally inefficient because of both the need to raise the funds to finance the subsidy and the inefficient product prices that result. Thus, voluntary approaches at the local level will generally have to rely on other participation incentives (e.g., appealing to local commitment to water quality improvement). Managers considering reliance on a voluntary approach should evaluate how likely it is that people would participate in the program, since this will be a key determinant of the effectiveness of a purely voluntary approach. Without these incentives, a purely voluntary approach may not provide sufficient protection. Managers may be able to increase the likely success of a voluntary approach by making it clear that, if the voluntary approach does not appear to be working, an approach based on regulation and/or financial penalties will be adopted. A mandatory approach based on regulations or taxes places a greater burden on the pollution sources, but if compliance can be ensured, it can be more effective in achieving water quality goals than a purely voluntary approach. However, when the costs of control vary across sources, uniform regulations will not meet those targets at the lowest possible cost. Cost-effective reduction can be achieved by allowing loading allocations to be traded. For example, allowing trades between point and nonpoint sources can generate significant cost reductions. In addition, managers can distribute initial permits in a variety of ways (e.g., uniformly across sources) without affecting the cost-effectiveness of the program. Marketable permit systems can, however, involve substantial administrative and information costs, and they may not work well if the number of sources that could participate in the permit market is small. The likely gains (in the form of cost reductions) must be weighed against the likely costs of using such a system. A careful examination of the effectiveness of trading in settings already employing it should be undertaken so managers have a better understanding of when this approach should be used. The choice among alternative policy instruments will depend on the nature of the available control options and the characteristics of the watershed. For example, for estuaries where a primary nutrient source is agricultural production within the watershed, managers can choose among policy instruments designed to reduce nutrient runoff, such as the provision of technical or financial assistance for the adoption of best management practices (see Chapter 9), regulations requiring adoption of those practices, mandatory soil testing, taxes on fertilizer use, and land use taxes. Alternatively, if a primary nutrient source is atmospheric deposition, local managers will need to work with regional or federal officials to develop strategies that reduce nutrient inputs to the estuary. While the
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution same types of policy options exist (e.g., technical or financial assistance for energy conservation or other reductions in NOx emissions, regulations limiting allowable emissions, energy use taxes, and emission taxes), these policies would generally have to be implemented regionally or nationally to combat atmospheric sources of nutrients. Because the appropriate choice of both the water quality target and the choice of policies to achieve that target are site-specific, a national recommendation regarding policy design is inappropriate. However, as part of a national strategy aimed at helping local managers reduce nutrient over-enrichment, the Committee recommends that a web-based clearinghouse for information relating to nutrient over-enrichment be developed. One component of that clearinghouse should be the compilation of three types of information that would aid local managers in developing nutrient management strategies that are appropriate for their estuaries: The first type of information would be a summary of and guide to research on the economic impacts of alternative source reduction methods, with particular emphasis on the role of site-specific characteristics in determining those impacts. This information would allow a local manager to determine which source reduction methods are likely to be more effective and cost-efficient, given the characteristics of the watershed and estuary of concern. For example, a manager of a local estuary with excessive nutrient inputs from agriculture could find information on the cost and effectiveness of various agricultural best management practices (see Chapter 9). The second type of information would be a summary of and guide to research on the effectiveness of alternative policies in achieving the most effective forms of source reduction, again given local circumstances. For example, if particular best management practices are identified as effective for a given watershed, this second type of information would provide a manager with information on the likely effectiveness of alternative policies in promoting increased adoption of those practices. Thirdly, the clearinghouse should contain documented case studies of both successful and unsuccessful attempts by local managers to combat nutrient over-enrichment in different types of estuaries. This would include not only attempts based on local policy implementation, but also documentation of attempts by local managers to work with regional or federal officials to combat nutrient loadings that originate outside the watershed, such as those from atmospheric sources. By providing meaningful and easily understood information about
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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution both the results of scientific research on source control and policy design, and on what has or has not worked in practice in different settings, local managers can increase their understanding of the likely effectiveness of alternative policies and hence make informed decisions about which policy approaches are most appropriate for them.
Representative terms from entire chapter: