Strategies for Meeting the Goals
The Chesapeake Bay estuary is one of the nation’s unique and valuable environmental resources. Preservation of this important ecosystem and proper evaluation and maintenance of its water quality are high priorities for the federal government, the Bay jurisdictions, and their citizens. As discussed in Chapter 1, the Chesapeake Bay Program (CBP) has established water quality goals for the Bay to address the adverse effects caused by nitrogen, phosphorus, and sediment loading from human activities and land development in the watershed. The primary sources of these pollutants include animal and crop agriculture, urban and suburban runoff, wastewater discharge via wastewater treatment plants and septic systems, and air pollution (see Chapter 1).
Bay jurisdictions have developed broad watershed implementation plans (WIPs) to implement practices by 2025 that will ultimately reduce nutrient and sediment loads by the amount necessary to attain the Bay water quality criteria. Reaching these goals will not be easy, however, and will require substantial commitment and, likely, some level of sacrifice from all who live and work in the Bay watershed. Jurisdictions will not only have to make significant reductions in current loads, but they will need to make additional cuts to address future growth and development over the next 15 years. Implementation strategies for the near-term have been developed for the first of the two-year year milestone periods, and detailed strategies through 2017 are in development through the Phase II WIP process. To reach the long-term load reduction goals, Bay jurisdictions and the federal government will need to consider a wide range of strategies, including some that are receiving little, if any, consideration today. Additionally, Bay part-
ners will need to adapt to future changes (e.g., climate change) that may impact the response of the Bay to reduced loads.
In this chapter, the committee takes a broad view of its task to discuss improvements in the development and implementation of strategies to meet the sediment and nutrient reduction goals (Task 6, see Box S-1). The chapter covers two main topics. First, the committee discusses future challenges in implementing effective remediation actions, including adapting to future changes in the drivers of Bay degradation and adapting to factors, such as climate change, which may alter the mechanisms of Bay recovery. Additionally, the committee discusses the costs associated with nutrient and sediment management actions and the challenge of maintaining political and public will. Second, the committee presents a range of strategies that could be used to help the CBP meet its restoration goals. These strategies encompass a wide range of topics, including practices, policies, funding strategies, and programmatic science management changes that have promise for improving the likelihood of attaining overall restoration goals.
Several cross-cutting issues could affect the pace and likelihood of achieving CBP goals. These challenges include expanding pressures on the Bay, such as population growth and development, changes in agriculture, and climate change. Additional challenges discussed in this section include costs and political and public will.
Shifting Drivers of Bay Water Quality and Ecosystem Response
The Chesapeake Bay’s ecological integrity and, hence, economic and social value has deteriorated because the ability to prevent excess nutrients and sediment from being discharged into the Bay has not kept pace with the generation of nutrients and sediment from rapid population growth and intensification of agricultural operations. These activities, combined with new economic challenges and impacts of climate change, will continue to challenge Bay restoration efforts. The success of an enhanced focus on water quality in the Bay will be, to a large extent, dependent upon the degree to which current plans (e.g., the total maximum daily load [TMDL]) and future efforts anticipate and respond to these challenges. This section of the report discusses trends in activities that are driving water quality problems in the Bay and the role that additional stressors may play in the ability of the CBP to meet future challenges.
Urban Issues: Population Growth and Development
In 2007, the U.S. Environmental Protection Agency (EPA) Inspector General concluded that new development in the Chesapeake Bay watershed had increased nutrient and sediment loads at rates faster than urban restoration efforts had reduced them (EPA OIG, 2007). This conclusion was reinforced by the CBP 2009 Bay Barometer report, which stated that pollution from urban and suburban areas continues to hinder the effectiveness of restoration efforts (CBP, 2010a). Phase 5.3 Watershed Model outputs estimate that total nitrogen loads from urban runoff and septic systems grew by 7.7 percent between 1985 and 2009; total phosphorus loads from urban runoff grew by 5.8 percent and sediment loads by 4.0 percent (see Appendix A). Urban and suburban sources of nutrients and sediment remain the only categories that continue to increase in modeled scenarios.
Population growth, development, and wastewater management combine to produce the observed impacts of urban and suburban development on water quality. The population of the Chesapeake Bay watershed grew from 8.1 million in 1950 to almost 16 million in 2000 (Claggett, 2007). Population growth estimates suggest that by 2030 the population will exceed 19 million (EPA OIG, 2007).
Distribution and patterns of population growth and development across the landscape have a major effect on water quality. Low-density, land-extensive residential development has combined with land-extensive development for other purposes (e.g., business, government), with connecting networks of impervious roadways and parking lots. More recently characterized as sprawl, this development pattern means that the rate at which open space is converted to support population growth outpaces population growth rates. Between 1990 and 2000, the watershed population increased by 8 percent, but the amount of land converted to development more than doubled. Based on projected population growth and the rate of growth in land development, the area of developed land could increase by more than 60 percent by 2030 (Boesch and Greer, 2003).
Sprawl development brings with it significant increases in the amount of impervious surface area, which channels water, nutrients, and sediment to waterways and minimizes the potential for landscapes to absorb them (Claggett, 2007). Between 1990 and 2000 impervious surface area in the Bay watershed increased by 41 percent (Claggett, 2007), and a 2006 study reported that impervious surface accounted for 18 percent of all urban lands in the Bay watershed (Tilley and Slonecker, 2006). Research suggests that stream water quality can be impaired when impervious cover in a watershed exceeds 5-6 percent (Couch and Hamilton, 2002).
Population growth and development patterns directly influence nutrient loading from wastewater. Wastewater treatment plants (WWTPs) collect
and treat wastewater from 75 percent of the households in the watershed, and technology upgrades have substantially reduced nutrient loadings from wastewater to the Bay (see Figures 1-12 and 1-13). However, private septic systems continue to present a challenge. The 2003 Chesapeake Futures report noted that approximately 25 percent of the housing units in the watershed were served by septic systems, contributing an estimated 33 million pounds of nitrogen per year to the watershed. Advanced nitrogen-removing septic designs exist, but they generally are not required for new development (Boesch and Greer, 2003). According to figures produced by the CBP, each new person added in homes built on septic systems results in about 3.6 pounds of nitrogen entering the local stream. By contrast, for homes connected to a state-of-the-art wastewater treatment plant, each new person adds only 1.6 pounds of nitrogen (Blankenship, 2006). For more than 50 years, residential development trends in the United States have been toward larger homes on larger lots at greater distance from urban centers with heavy reliance on septic systems for wastewater management.
If population, land development, and reliance on private septic systems continue to grow, the challenges of reducing nutrients and sediment entering the Bay will continue to grow. Simply managing development that comes with population growth may not be sufficient to meet water quality goals. Tom Horton (conservationist) has argued that attention to restricting population growth may be needed:
Our environmental impacts are the sum of how many of us there are, and how much each of us demands of the air, water and land. That is our total environmental ‘footprint’. Common sense tells us we can help the Chesapeake Bay and the planet by reducing either per capita impacts, or the number of capitas. It also tells us that if one side of the footprint equation keeps increasing, we will gain that much less from just working the other side (Horton, 2008).
Agricultural Issues: Changes in Animal and Crop Agriculture
Agriculture is an integral component of the culture, heritage, and economy of the Bay watershed, and as of 2003, agriculture accounted for 13 percent of the region’s gross domestic product (GDP). However, agriculture’s share of GDP has steadily declined over the past decades (Boesch and Greer, 2003). Between 2002 and 2007, cropland and farmland acres declined by 10 percent and almost 15 percent, respectively. Furthermore, the type of agriculture being practiced is shifting. Even though the total amount of nutrients and sediment entering the Bay and its tributaries from agricultural sources has decreased since 1985, the agriculture sector has been responsible for a smaller portion of reductions than have point sources, especially municipal wastewater treatment plants. This imbalance
can be explained by the fact that reduction efforts by agriculture tend to be voluntary and incentive-based, whereas efforts by point sources are dictated by regulation.
Changes in Animal Agriculture. In 2007, there were 16.8 million people living in the Bay Watershed along with 2.4 million cattle, 1.2 million hogs, and 222 million chickens (U.S. Census of Agriculture, 2007).1 The waste generated by these populations contributed 40 percent of the total nitrogen load to the Bay watershed: 23 percent of the total from human sewage (i.e., septic systems and municipal and industrial wastewater) and 17 percent from animal manure (see Figure 1-6). Although projections for future changes in human population are readily available,2 projections for changes in the animal population are not. However, based on current trends, the number of animal production operations (including dairy) is predicted to decrease as a result of continuing industry consolidation within the Bay watershed. Yet, the number of animals per operation is predicted to continue to increase to meet growing demand, especially from a growing regional market (Mark Dubin, University of Maryland, personal communication, 2010). With fewer but larger operations, the total number of animals may well be maintained or even increased. However, the species mix may change. Over the period of 2002 to 2007, cattle, sheep, and swine populations decreased, while the numbers of chickens, horses, and goats increased. Of these, the most notable was the increase in chickens. At this point, formal projections for changes in animal populations are not available from the CBP (Mark Dubin, University of Maryland, personal communication, 2010). If the trend towards increased livestock concentration continues, more animal production operations will be classified as point sources, which will bring them under NPDES regulatory requirements and presumably reduce contributions to nutrient loads.
Agricultural Production and Land-Use Changes. Shifts in agricultural production and land use often occur because of external pressures, and these changes have implications for nutrient management in the Bay watershed. For instance, the drive for biofuel production to provide a greater share of consumed energy, often required by law,3 could lead to increased nutrient loading from agricultural lands. Between 2005 and 2010, corn acreage in
1 Note that the 2007 Census of Agriculture numbers reported are for mid-atlantic subwatersheds that drain to the Bay. The human population is from the CBP, available at http://www.chesapeakebay.net/status_population.aspx?menuitem=19794.
3 For example, the American Recovery and Reinvestment Act of 2009 required 12.5 billion gallons of biofuel (primarily ethanol) be mixed with gasoline by 2012.
Bay watershed states increased by 11 percent, mostly on land removed from soybean production, the Conservation Reserve Program, and pastureland (USDA National Agricultural Statistics Service, 2010). The potential for nitrogen and phosphorus loss from corn production is greater than most other land uses (see Box 5-1).
Additionally, the incorporation of dry distiller’s grain (DDG), a by-product of ethanol production, into beef and dairy cattle rations could erode progress in managing nitrogen and particularly phosphorus in animal feed (to reduce nutrients in manure). Using DDGs as a feed ration alternative is likely to increase because of its ready availability and low cost relative to corn grain prices. However, the phosphorus content of DDGs (0.8-0.9 percent phosphorus) is about three times that of corn, which makes it dif-
Shifting Nutrient Loads from Agricultural Land Use Changes
Corn is an inherently inefficient nitrogen user; 40 to 60 percent of nitrogen applied generally is not taken up by the crop, and nitrogen loads to downstream aquatic ecosystems from corn-dominated landscapes are typically 25 to 45 lb nitrogen ac-1 yr-1 (Balkcom et al., 2003; Randall et al., 2003). Nitrogen losses to aquatic systems from soybeans average 18-35 lbs nitrogen ac-1 yr-1 (CBP, 2006). Similarly, average phosphorus losses in runoff from corn (3-18 lbs ac-1 yr-1) tend to be greater thanfrom soybeans (1-10 lbs ac-1 yr1) (Carpenter et al., 1998; Kimmell et al., 2001; Sharpley and Rekolainen, 1997). The loss of phosphorus from perennials and hay crops (0.2-1 lb ac-1 yr-1) is generally less than from annuals because runoff volumes are lower and crop phosphorus requirements are smaller, so smaller amounts of fertilizer or manure are applied (Sharpley et al., 2001; Smith et al., 1992). Further, water-quality model simulations of converting Conservation Reserve Program acreage or perennial grasses to cropland confirm that delivered nitrogen and phosphorus loads increase by more than double the percentage land area converted (Mankin et al., 1999, 2003). Assuming fertilizer application rates remained constant, the estimated 0.25 million acre increase in corn acreage (0.1 million ha) over the past five years in the Chesapeake Bay Watershed is projected to have increased annual nutrient loads by 5 million lbs nitrogen and 2 million lbs phosphorus (Table 5.1).
ficult to use such materials at more than 15 percent of animal feed rations without exceeding dietary phosphorus recommendations (Lawrence, 2006; NRC, 2000). The inclusion of DDGs in rations exceeding recommended rates will increase the phosphorus content of manure (Baxter et al., 2003; Maguire et al., 2004; Wu et al., 2001) and, if the manure is land-applied, increase the potential for phosphorus loss in runoff (Ebeling et al., 2002; Maguire et al., 2007; Sharpley et al., 2005).
Climate change is likely to affect the Bay’s response to nutrient and sediment management controls. However, uncertainty exists in predicting
TABLE 5-1 Estimated Increase in Nutrient Export in Farm Runoff from Growing an Additional 0.25 Million Acres of Corn in the Chesapeake Bay Watershed
|Acreage Shift to Support Ethanol||Land Area
|Nitrogen Export in Runoff||Phosphorus Export in Runoff|
New corn acres
Converted from soybeansa
Converted from CRP landa
Converted from idle, pasture or hay landa
Estimated increased nutrient export in runoffb
aNutrient export of nitrogen and phosphorus if land had remained in soybeans, Conservation Reserve Program, idle, pasture, or hay land.
bIncrease in nitrogen and phosphorus export in runoff estimated as that occurring from additional corn acres minus the runoff that would have occurred from the original land use prior to conversion to corn.
SOURCE: Adapted from Simpson et al. (2008).
Historical Climate Changes and the Effect on Hypoxia
The relationship between spring nitrate loading from the Susquehanna River (a proxy for total nitrogen loading to the Bay) and the resulting summer time anoxic water was examined by Hagy et al. (2004) and is shown in Figure 5-1. There appears to be a change in the anoxia’s response to nitrogen loading. The anoxia that developed during the latter years (1980-2001) was significantly greater than that which developed during the early years (1950-1979) for the same winter-spring loading. For example, a January-May loading of 20 gigagrams (Gg), resulted in an anoxic volume of approximately 1 km3 in the early years and perhaps 3 km3 in the latter years. There is significant scatter in these data, and the loading and anoxic volume estimates in the early years are less reliable than those subsequent to 1985 when more intensive monitoring became available.
This striking result has generated a number of hypotheses, including the idea that a “regime shift” has occurred in the biology (Petersen et al., 2008). More mechanistic hypotheses have been advanced, but all are associated with changes that are related to climate variation. Suspected mechanisms include wind direction (Scully, 2010a,b); increase in water temperature, decrease in oyster abundance and associated filtration capacity, and less efficient nitrification-denitrification (Kemp et al., 2009); and decreased mixing of waters and increases in early summer stratification (Murphy et al., 2011). In particular, Murphy et al. (2011) have suggested that the decreased mixing of waters and increased early-summer stratification may relate to both an observed shift in the predominant wind direction over the Bay (Scully, 2010a) and to observed increases of Bay salinity levels, which have in turn been related to sea level rise (Hilton et al., 2008).
climate change effects on the forcing functions to the Bay (e.g., magnitudes and timing of rainfall and runoff, range and patterns of temperature variation, influence of storm activity) and how the Bay’s physical, chemical, and biological systems will respond. Nevertheless, attempts have been made to quantify the possibilities, using models and professional judgment.
Najjar et al. (2010) published a comprehensive examination of the potential responses of the Bay to climate change and concluded that “likely changes” include increases in precipitation amount and intensity, salinity variability, harmful algae, hypoxia, and coastal flooding. Annual mean temperatures in the Bay Watershed are projected to increase by 1°C during
FIGURE 5-1 Midsummer volume of anoxic bottom water vs. winter-spring nitrate loading from Susquehanna River for earlier years (1950-1979, solid line, filled circles) and for later years (1980-2001, dashed line, open circles).
SOURCE: Kemp et al. (2005), modified from Hagy et al. (2004).
the next 30 years and perhaps by 7 °C by the end of this century. Najjar et al. (2010) concluded: “Climate change alone will cause the Bay to function very differently in the future.” Such changes include altered interactions among trophic levels and a reduction in eelgrass, the dominant species of underwater grasses in the Bay. Even small changes in water temperature are projected to have significant impacts on fishery resources. Climate changes during the past 50 years also appear to be affecting the extent of Bay hypoxia that can be detected in observational data (Box 5-2). Climatic variations can dramatically change the Bay’s response to nutrient management actions. To avert overly optimistic expectations, and the associated
disappointment from achieving less than forecasted results, a systematic investigation should be initiated. Coupling the currently available models, validated by hindcasting historical data, with the available climate change scenarios would be a first step.
Meeting the overall cost of Bay management is a key challenge facing the CBP. In October 2004, the CBP’s Blue Ribbon Finance Panel considered the entire 64,000 square mile watershed and estimated total water quality restoration costs at $28 billion (or $32 billion in 2010 dollars).4 This is equivalent to approximately $1,900 (in 2010 dollars) for each of the 16.8 million residents in the Bay watershed. The recent release of the watershed implementation plans (WIPs) has generated additional concerns about the costs of implementation (EPA, 2010f). For example, Maryland estimates costs to residents, businesses, and taxpayers of meeting the goals of its Phase I WIP at $13-15 billion (MDE et al., 2010). Virginia estimates costs in that state of implementing its Phase I WIP will exceed $7 billion (VA DNR, 2010). Undoubtedly, the costs reported in these and other documents are rough estimates at best, but their magnitude is indicative of the financial challenges posed.
Costs beyond those included in financial calculations are also possible. For example, restrictions on land-use changes and limits on growth could be manifested over time in higher housing costs. Efforts to reduce airborne nutrient sources could raise the cost of energy generation or transportation. Lifestyle changes also may be required, such as restrictions on residential and commercial landscaping (e.g., restricted fertilizer applications) and greater reliance on public transportation. However, such lifestyle changes could offer economic benefits, such as reduced day-to-day cost of living for individuals and reduced emissions of greenhouse gases. Also, many BMPs may offer broader benefits than just those targeted; for example, reducing stormwater runoff volume to protect water quality may reduce flood damages.
The costs of nutrient and sediment reduction are unlikely to be evenly distributed. Even if nutrient and sediment reductions are distributed across the Bay jurisdictions based upon their relative contributions, the shares of the overall cost borne by any jurisdiction’s residents, businesses, and taxpayers are likely to vary depending upon the specific sources within the jurisdiction’s boundaries and the location of the jurisdiction within the watershed. Much of the public cost will be absorbed by taxpayers as local governments deal with upgrades to wastewater treatment plants and
stormwater management, and higher taxes and fees are always contentious. Distribution of costs across regulated and unregulated sources will differ. All states have indicated that they will rely upon federal funds to cover a substantial portion of implementation costs. In light of other federal budgetary pressures, sufficient federal assistance is unlikely.
Costs to Agriculture
The costs to agriculture of implementing BMPs to further reduce nutrients and sediments will be a function of the mix of land management requirements adopted by states, financial assistance from federal, state, and local governments, and the financial benefits that may offset some or all of the costs incurred by farmers’ groups. A substantial body of research indicates that many agricultural practices that reduce nutrient and sediment loss from farms can offer economic benefits and sometimes provide competitive gains at the farm level (NRC, 2010). However, some practices may impose significant on-farm costs that are not recovered. Agricultural producers are not able to set the prices they receive for their products and, as a result, are not able to make adjustments to cover additional costs of implementing BMPs. Farmers groups have questioned whether agriculture can bear the costs of additional BMPs (American Farm Bureau, 2010).
Estimates of costs to agriculture vary widely, and how the relative on-farm costs and benefits of individual BMPs are accounted for in these estimates is not clear. In 2005, the CBP estimated the cost of implementing agricultural portions of Bay cleanup strategies at about $700 million a year, with only $188 million in conservation funding provided each year by the 2008 Farm Bill (Blankenship, 2008). More recent estimates suggest that the costs to agriculture could be even higher, and the future availability of federal and state subsidies is uncertain. Without doubt, the costs of reducing agricultural sources of nutrients and sediment will be very high. Deciding how and among whom the costs will be distributed represents a substantial challenge.
Costs in Urban Areas
Efforts to control nutrients and sediments in the Bay watershed have had, and will continue to have, significant effects on the way municipalities and industries manage their land, wastewater, development, and redevelopment. The seven Bay jurisdictions have identified ambitious plans necessary to meeting the wasteload and load allocations required by the TMDL (EPA, 2010a). Although they do not fully articulate the potential sector-specific costs associated with TMDL compliance, together with information from the 2008 Clean Watersheds Needs Survey (EPA, 2010g) they provide
enough information to indicate that costs to enhance wastewater facilities and control stormwater and nonpoint source runoff from urban areas will be in the billions of dollars. The Bay jurisdictions cite nearly $70 billion in existing wastewater and stormwater management needs (including state needs for facilities located outside the Bay watershed) to meet the goals of the Clean Water Act (EPA, 2010g). For the five jurisdictions reporting this information, meeting the 2011 milestone goals, alone, will require federal, state and local funds totaling more than $2 billion to address urban point and nonpoint sources (CBP, 2009b).
The Chesapeake Bay WIPs identified costs of urban stormwater BMPs (including retrofitting and low impact design [LID]) ranging from just a few thousand dollars per impervious cover acre treated to as much as $200,000 per acre. Implementing stormwater BMPs on existing development is typically costly because of constraints such as the lack of space to install BMPs and the often-prohibitive cost of purchasing land for BMP installation. (See Box 5-3 for an example.) By contrast, installing stormwater management practices at the time of new development is more cost-effective, and with proper planning LID may even include some cost savings over conventional development (Schueler et al., 2007; Schueler, 2009). With public
A Connecticut Example of Urban
Stormwater Management Costs
In 2007, Connecticut issued the first TMDL based on impervious cover that provides a surrogate approach for addressing a multitude of pollution problems related to development, including nutrients and suspended solids (CT DEP, 2007). The TMDL impaired watershed is dominated by the University of Connecticut campus, which provided an excellent opportunity, primarily within one land ownership entity, to develop a management plan that incorporates state-of-the-art LID retrofit technologies (Center for Watershed Protection and Horsley Witten Group, 2010). The analysis determined that actions could be taken to mitigate the effect of impervious cover and meet water quality standards, but at a substantial cost. High priority (“top 10”) actions would address 30 acres of impervious cover at a cost of $1.35 million, or about $45,000 per acre of impervious cover. A full retrofit scenario addressing nearly 61 acres of impervious cover would cost $5.8 million or about $95,000 per acre of impervious cover. Both are in the range of costs estimated in the WIPs and are substantial.
acceptance, some LID practices such as pervious pavers, naturalistic landscaping, and broader open space can add value to a development. In any case, requiring all developers to abide by the same restrictions prevents any imposition of competitive disadvantage, a primary concern of many (Fuss and O’Neill, 2010a).
Public Support and Political Will for Attaining Goals
Despite the importance that watershed residents place upon Bay water quality and all of the ecosystem and economic benefits that are expected to result from improved water quality, questions remain about watershed residents’ willingness to bear the expected costs of reducing nutrient and sediment loads to the Bay and its tributaries. For one thing, the benefits of reducing nutrient and sediment loads are not evenly distributed across the watershed. Of the seven jurisdictions that lie within the watershed, only Maryland and Virginia directly border the Bay’s mainstem. As a result, residents of the other states may question why they should finance programs designed to help water quality in the Bay itself.
Residents of Delaware, New York, Pennsylvania, West Virginia, and the District of Columbia likely are more concerned about water quality within their own boundaries than for downstream water bodies including the Chesapeake Bay. New York, Pennsylvania, and West Virginia also have significant land area that falls outside the Bay Watershed, which means that water quality programs in those states must balance efforts between the Bay watershed and other watersheds.
Residents’ willingness to invest in environmental protection and resource conservation is demonstrated by recent resident-supported bond initiatives in New York, Pennsylvania, and Virginia. New York’s 1996 initiative, which passed by 57 to 43 percent, provides for bonds to fund clean water and clean air programs. Pennsylvania voters passed bond initiatives in 2004 (63 to 37 percent) to support water and wastewater infrastructure improvements, in 2005 (61 to 39 percent) to support land conservation, and in 2008 (62 to 38 percent) to support water and sewer improvements. Virginia voters approved bond initiatives in 1992 (67 to 33 percent) and 2002 (69 to 31 percent) to support investments in parks and recreational facilities, including land acquisition and capital improvements.5
Likewise, state legislatures have made funds available for nutrient reduction programs. For example, Maryland’s legislature voted in 2004 to levy fees on each household served by a wastewater treatment plant and each household served by an on-site septic system; the funds are used to
5 Election results from the Trust for Public Land’s LandVote database are available at http://www.tpl.org/tier3_cd.cfm?content_item_id=12010&folder_id=2386.
support upgrades for wastewater treatment facilities and septic systems in order to reduce nutrient discharges (Marx, 2004). Virginia’s legislature voted in 2007 to sell bonds to fund nutrient removal technologies at wastewater treatment plants (Code of Virginia § 10.1-1186.01.).
Whether public support for additional investments in Bay improvements will continue is an open question. State and federal economic climates have forced citizens and their elected officials into recurrent rounds of fiscal belt-tightening. High unemployment combined with tight state budgets may well leave state residents focused on priorities other than water quality in the Chesapeake Bay in particular and environmental concerns generally. Historically, public concerns about environmental quality have not always been accompanied by public willingness to bear the costs of achieving environmental improvement and protection (Gillroy and Shapiro, 1986), especially when economic health is tenuous (Franzen and Meyer, 2010). However, economic and environmental initiatives can be coupled. The American Recovery and Reinvestment Act of 2009 provided $105 billion for infrastructure improvements, including $4 billion nationally for Clean Water State Revolving Funds and an additional $2 billion for Drinking Water State Revolving Funds.
The CBP’s success hangs on political will as much as individual citizen willingness to shoulder the financial burden. Questions about whether such political will exists are widespread (Pegg, 2004; Thompson, 2004; Ernst, 2006; The Monitor’s View, 2009; Wood, 2010). Political scientist and author Howard Ernst has described the “political dead zone,” in which elected officials profess concern about the Bay but fail to make the hard decisions necessary to achieve real improvement. Polluting industries continue to pollute, and the environmental community lacks the influence needed to advocate successfully for the Bay (Ernst, 2009).
Elected officials are generally hesitant to implement environmental policies that are perceived to challenge economic development, even if the policies promise environmental improvement. When specific business sectors or communities expect that costs of environmental policies will affect them disproportionately, they see a direct benefit from lobbying elected officials and voicing their opposition. Challenges to the Chesapeake Bay TMDL have been heated (e.g., Agricultural Nutrient Policy Council, 2010; American Farm Bureau, 2010; ESA Policy News, 2010; Harper, 2010; Stuart, 2010). In contrast, environmental benefits will be diffuse and enjoyed by a broad cross-section of the population. Such beneficiaries are less likely to become directly involved in lobbying efforts because they know they will benefit from environmental policies even if they do not actively lobby for them. The EPA received more than 14,000 comments on the draft TMDL. More than 13,000 of those comments originated from mass mail campaigns
organized by more than 20 environmental groups. Of the remainder, the most detailed individual responses were from those voicing opposition to or some degree of reservation about the plan (Blankenship, 2010).
Compliance with the water quality criteria (e.g., for dissolved oxygen) is the measuring stick for success of the TMDL, but Bay area citizens may require more visible, tangible evidence of water quality improvement. Because of lag times between the implementation of land-based nutrient and sediment control practices and improvement in Chesapeake Bay water quality (see Box 1-3), many BMPs undertaken for the first milestone may not result in observable improvements in the Bay until much later. Likewise, the research and regulatory communities working on Bay issues may be aware of the uncertainties inherent in the load projections and Bay system dynamics that could result in outcomes that differ from those anticipated (see Chapter 4). However, absent sufficient articulation and explanation of those uncertainties, members of the public less schooled in the scientific elements may fail to understand or accept less than absolute, observable improvements. If these lags and uncertainties aren’t adequately explained, CBP partners will need to anticipate and be prepared to respond to the potential ramifications of an impatient or disillusioned public. Concerns that non-experts will be unable or unwilling to adjust their expectations based on a better understanding of uncertainties or time lags are likely unfounded. Propst et al. (2000) presented research evidence that individuals who lack knowledge of technical scientific issues can quickly learn about their critical features and choose policy options similar to those chosen by scientists and are likely to ask the right questions and find novel solutions.
If public and political support wanes in the face of high costs and time lags between BMP implementation and water quality improvement, a jurisdiction could formally question the feasibility of meeting water quality goals by requesting a use attainability analysis (UAA). A UAA is an assessment of the factors affecting the likely attainment of designated uses of a water body. A jurisdiction may request that a designated use be changed if it can be shown, through a UAA, that current designated use is precluded by physical, chemical or biological factors or that the stringency of controls needed would result in “substantial and widespread economic and social impact” (EPA, 2010i). Changing the designated use would result in the establishment of new water quality criteria, which could reduce both the efforts required to meet water quality goals and the costs of those efforts. Thus, if a jurisdiction is concerned about its ability to garner financial resources needed to address costs resulting from federal requirements for compliance with water quality criteria, it could assert public and political will by working to change designated use and reduce pollution control pressures.
STRATEGIES FOR IMPROVEMENT
Previous sections of this chapter presented a number of challenges the CBP could face in its efforts toward the CBP’s long-term nutrient and sediment reduction goals. In this section, the committee identifies strategies that could be used to help the CBP meets its goals. The committee did not attempt to identify every possible strategy that could be implemented but instead focused on approaches that are not being implemented to their full potential, or in some cases have substantial potential but are not being widely discussed. This section includes practices and policies for reducing agricultural and urban pollutant loads and air pollution strategies followed by a discussion of possible funding strategies. Because many of these strategies have policy or societal implications, the strategies are not prioritized but are offered to encourage further consideration and exploration among the CBP partners and stakeholders. Finally, the committee discusses the importance of modeling and monitoring to help meet the goals and recommends the formation of a Chesapeake Bay modeling laboratory as a strategy to improve the scientific and modeling support for the CBP.
Several strategies exist to improve nutrient management on agricultural land. Key strategies that could be used in the Bay watershed include, but are not limited to, improving management of animal agriculture and manure use and developing new incentive- and regulatory-based programs targeted at improved agricultural nutrient management.
Strategies for Improved Animal Agriculture and Manure Management
Intensive animal feeding operations (i.e., dairy, poultry, swine) are common locally significant sources of nutrient enrichment to surface water and groundwater. The potential for phosphorus and nitrogen surplus on farms dominated by concentrated animal feeding operations (CAFOs)6 can be much greater than on farms with cropping systems or integrated crop and
6 An animal feeding operation (AFO) is a facility that confines animals for more than 45 days in an area that does not produce vegetation during the growing season. CAFOs are AFOs that meet specific size and surface water discharge criteria or that have been designated on a case-by-case basis as significant contributors of pollutants by the state or local permitting authority (see http://www.epa.gov/npdes/pubs/sector_table.pdf, and http://www.epa.gov/region07/water/cafo/).
livestock operations.7 Nutrient inputs to CAFOs are dominated by feed, and often nutrients in feed exceed the amount that can cycle through a feed-crop production system for removal from the farm. Research has shown that only about 30 percent of nitrogen and phosphorus in feed is utilized by the animal, with the remaining excreted in manure (Poulson, 2000; Valk et al., 2000). As a result, in most CAFOs, animal feed is the primary source of on-farm nutrient excess.
Based on U.S. agricultural census surveys of livestock numbers over the past 15 years, a steady increase in the amounts of nitrogen and phosphorus accumulating in livestock operations exceeding crop requirements at the farm level has been described (Kellogg et al., 2000). Unless crop and livestock operations are combined or enter into contractual arrangements to optimize nutrient management, innovative approaches—beyond typical nonpoint BMPs expected of agriculture—will be needed to reduce or eliminate the nutrient imbalances. In this section, strategies for animal nutrition and manure management are discussed separately.
Animal Nutrition Management. Sustainable nutrient management in animal agriculture begins with sound feed decisions (i.e., nutrient concentrations in animal feed that match dietary recommended levels). Ebeling et al. (2002) showed that increasing the phosphorus concentration in the diet of dairy cattle doubled the potential for phosphorus export in runoff from land-applied manures, even with similar overall phosphorus application rates. This difference was likely due to a greater proportion of manure phosphorus being water soluble in manure with a high-phosphorus diet compared to a low-phosphorus diet. Similar trends have been observed in beef cattle, pigs, and poultry (Kleinman et al., 2002, 2005). Implementing a carefully planned diet tailored to meet the specific nitrogen and phosphorus requirements of animals in each phase of their growth will minimize nutrient loss to the environment in feces, urine, and gases. Reducing nitrogen and phosphorus in animal feed presents a promising nutrient management opportunity that can effect lasting reductions in nitrogen and phosphorus loads to the environment.
A reduction in the quantity of nitrogen and phosphorus excreted by livestock can also be accomplished by supplementing livestock diets with enzymes to enhance digestion (Keshavarz and Austic, 2004; Knowlton et al., 2002). Enzymes, such as phytase, can be added to feed to increase the efficiency of grain phosphorus absorption by pigs and poultry. Such
7 Annual surpluses of nitrogen and phosphorus were reported for grain-poultry farms in Delaware (Sims, 1997) and grain-dairy farms in New York (Klausner et al., 1998). These surpluses can vary from 105 to 1170 lbs N ac-1 yr-1 and from 23 to 1000 lbs P ac-1 yr-1 (Bacon et al., 1990; Lanyon, 2000).
enzymes reduce the need for phosphorus supplements in feed and potentially reduce the total phosphorus content in manure. Corn hybrids are also available that contain a lower percentage of indigestible phosphorus so that phosphorus availability to nonruminants is two to three times higher than from normal corn. Pigs and chickens fed “low-phytic-acid” corn excreted less phosphorus in manure than those fed conventional corn varieties (Ertl et al., 1998).
There are many opportunities through the use of conventional breeding and genetic engineering to improve the digestibility, nutrient utilization, and need for mineral additives in feed, as well as the digestive process of the animal (Abberton et al., 2008; McSweeney et al., 1999; Tabe et al., 1993). For example, a genetically enhanced line of Yorkshire pigs has been developed with the capability of digesting plant phosphorus more efficiently than conventional Yorkshire pigs (Forsberg et al., 2003). The salivary glands of these pigs produce the enzyme phytase, which in the acidic environment of the stomach, degrades indigestible phytate in the feed that accounts for 50 to 75 percent of grain phosphorus. Thus, there is no need to supplement the diet with either mineral phosphate or commercially produced phytase, and there is less phosphorus in the manure. Nevertheless, the public’s acceptance of and willingness to use genetically modified products remains to be seen.
CAFO and Manure Management. Only Maryland, Delaware, and New York currently require specific comprehensive nutrient management plans as a part of statewide CAFO permitting programs. Making these programs consistent across all states and including smaller animal production operations, which in aggregation can be locally significant sources of nutrients to surface and groundwater, would enhance nutrient management in the Bay watershed. Interestingly, the TMDL backstop allocations proposed for CAFOs throughout the watershed appear to assume that all animal feeding operations, regardless of size, are subject to the same nutrient management requirements as CAFOs. Changes in CAFO requirements at a national, rather than regional, level would avoid putting producers in the Bay watershed at a competitive disadvantage.
Manure is a valuable resource for improving soil structure and increasing vegetative cover, thereby reducing runoff and erosion potential (Risse et al., 2006). However, manures have historically been applied at rates designed to meet crop nitrogen requirements, providing at least twice as much phosphorus as crops need and resulting in the buildup of soil phosphorus above levels required for crop production. EPA (2010e) recommends that manure or fertilizer not be applied to any soil in the Chesapeake Bay watershed with a “phosphorus saturation” value above 20 percent. Manure applications in excess of crop nitrogen or phosphorus needs, causes
a concomitant increase in the potential for nitrogen and phosphorus loss via surface and groundwater flows and nitrogen gas emissions within the Bay Watershed (Kovzelove et al., 2010; Sutton and Cox, 2010). In the case of nitrogen emissions to the atmosphere, not only is N2O a global concern, but the ammonia and NOx emissions are a concern for nitrogen deposition back to the Bay and its watershed, and to downwind regions.
Transport of manure from nutrient-surplus to nutrient-deficit areas can address imbalances as long as the nutrients are appropriately applied on receiving lands to avoid potential losses to water and air. Manure transport out of the Bay watershed is of long-term benefit and is occurring through subsidized initiatives of poultry integrators, with poultry litter that is dried, ground, and compacted into small, less bulky pellets (pelletized litter). However, wet and heavy manures are not being transported more than a few miles from where they are produced because of cost and technical difficulties. Wider adoption of manure transport that links producers with buyers of manures for crop fertilization will greatly enhance the sustainability of animal operations over a larger geographic area. In attempts to address this, the Natural Resources Conservation Service (NRCS) has developed a cost-share program that could facilitate manure transport from surplus to deficit areas. As the costs of implementing more complex or restrictive conservation or remedial measures increase, transport will become more economically viable. Additionally, as energy prices increase, alternate uses for manure, such as burning for electricity generation and digestion for methane production, will become more attractive. However, with bioenergy production, the nutrient-rich biochar (residues remaining after burning) or sludge (solids remaining after digestion) still need to be managed appropriately.
Another way to significantly decrease nutrients inputs to the Bay would be to limit the extent of animal operations to the nutrient carrying capacity of the watershed, considering the existing loads. Under such a scenario, if existing operations were unable to reduce their nutrient loadings through innovative manure management, some percentage of the animal protein production would need to be outsourced from the watershed. Clearly, this alternative has negative socioeconomic consequences (e.g., employment losses, economic decline of agricultural and rural communities), but it is an example of the type of bold action (and difficult policy decision) that may be needed to restore the Bay’s ecosystem.
Incentive-Based BMP Programs
Voluntary incentive-based BMP programs can provide a low-cost approach to improve BMP implementation, maintenance, and tracking. Key to their success is identifying an incentive that has value to the landowner.
Providing landowners with regulatory relief has proven effective in Florida (see Box 2-2). A presumption of compliance with water quality standards given in exchange for voluntary BMP implementation, maintenance, and reporting has proven to be a powerful incentive. In Florida, participants of the incentive-based BMP program are required to maintain nutrient management plans and provide access to land management records and land parcel identification. See Chapter 2 for a more in-depth discussion of incentive-based BMP programs. USDA has recently begun discussions with EPA and Bay jurisdictions about creating a similar program in the Bay watershed, where farmers could agree to implement certain practices in exchange for presumptive compliance with regulations (A. Mills, USDA, personal communication, 2011).
European countries have taken a regulatory approach to reduce agricultural nutrient contamination, and these strategies offer another model for possible consideration by the CBP. In many respects, agriculture in the Bay watershed is similar to agriculture in Denmark, where trends toward larger and more specialized farms with high animal production in the western parts of the country are spatially separated from specialized crop production in the eastern parts of the country. Nutrient management legislation was enacted in Denmark in the 1980s as part of the European Union’s “Environmental Action Plan” and “Water Framework Directive” (De Clercq and Sinabell, 2001). See Box 5-4 for details of Denmark’s agricultural nutrient management regulations.
Coupling regulatory requirements with incentives in Denmark created agricultural production systems that manage nutrients to limit surplus. These policy changes resulted in a decline in the national nitrogen surplus from 132 lbs nitrogen ac-1 in 1980 to 79 lbs nitrogen ac-1 in 2006 (a 41 percent decrease) (Kyllingsbæk and Hansen, 2007). Over the same period, phosphorus surpluses decreased from 26 to 10 lbs ac-1 (a 62 percent decrease), with a concomitant increase in the crop uptake of phosphorus from 24 percent in 1980 to 56 percent in 2006 (Maguire et al., 2009).
In Denmark, between 1989 and 2002, a significant decrease in total nitrogen concentrations occurred in 48 streams draining agricultural watersheds without major point sources as a result of these regulatory and incentive-based measures (Kronvang et al., 2005). The downward trend became more evident as the proportion of agricultural land in the watershed increased and was more pronounced for loamy than for sandy soils. In contrast, no significant trends could be detected for total phosphorus concentrations in streams draining agricultural watersheds. As in the Chesa-
Denmark’s Agricultural Regulations to Enhance Water Quality
Danish agricultural regulations require specific measures (De Clercq and Sinabell, 2001), including:
• Nitrogen fertilizer applications are limited to 90 percent of the optimum for each crop.
• Manure can be spread on land from the beginning of February until harvest time in August. Manure can be spread until the end of September on certain crops such as grass and oilseed rape.
• At least 70 percent of the nitrogen applied in manure must be recovered by plant uptake.
• Manure storage capacity for a minimum of 9 months must be provided,with cost share incentives to construct storage facilities.
Additionally, almost all farmers are required to produce a yearly nutrient management plan. The only farmers that are exempt are those with fewer than 20 animal units (e.g., 20 mature cows under 1,000 lbs, 20 horses, 66 pigs weighing between 50 and 300 lbs, and 6,667 broilers under 5 lbs) and a low stocking rate (few animals per acre), often termed “hobby farmers.” However, these farmers pay a tax on the nitrogen that they purchase.
Even though a surplus of phosphorus can still be added in areas with the highest animal densities, an upper limit of the phosphorus surplus was indirectly introduced with these regulations. Furthermore, the requirement for 9-month storage capacity for animal manure made it possible to shift from autumn to spring application. This resulted in further improvements in nutrient utilization and likely resulted in a reduction in manure-related phosphorus loading during the wet winter season (Schelde et al., 2006).
peake Bay Watershed, a lack of a response in phosphorus concentrations to management measures was attributed to legacy phosphorus and its resilience in lakes and estuaries, which may delay the full effects of such action plans (Maguire et al., 2009).
Even though the water quality results in Denmark are compelling, applying a similar such program with enhanced regulatory oversight of agriculture in the Bay watershed would be politically challenging in light of recent opposition to further regulation of agriculture by EPA voiced by
several agricultural groups (Copeland, 2010). Further, farmers in the Bay watershed would be placed at a competitive disadvantage compared with those outside the watershed unless such changes were imposed nationally or unless financial support programs were put in place. One way to avoid such a competitive disadvantage is to establish a set of minimum performance standards that apply nationally. For example, Cox (2010) has described the role of what he calls “precision regulation” that would regulate behavior that is unequivocally damaging to water quality, such as requiring that cattle be fenced away from water bodies, that manure not be spread on frozen ground, or that there be riparian setbacks of crops from the edges of water bodies (Batie, 2009).
As noted in Chapter 1, the Watershed Model estimates that total nutrient and sediment loads from urban runoff and septic systems increased between 1985 and 2009, while loads from all other sectors decreased (see also Appendix A). This is likely due to a combination of urban growth and insufficient levels of BMP implementation or performance. For the entire watershed, an additional reduction in nitrogen loads of 23 percent for urban stormwater and 24 percent for septic systems will be required to meet TMDL goals, an ambitious and costly reduction target considering the added pressures of continued population growth and development. Several potential strategies for addressing these challenges are discussed below, including regulatory approaches for stormwater management and fertilizer use, enhanced wastewater management, and encouraging increased individual responsibility.
Three classes of regulatory strategies that could be used to improve urban nutrient and sediment management address: 1) stormwater management, 2) offsets for growth and development, and 3) limits on residential fertilizer use.
Stormwater Management. Common municipal policies to promote improved stormwater management include local regulations, codes, and ordinances; stormwater incentives or fees (discussed later in the chapter); and education and outreach (EPA, 2010j). Primary regulatory tools for implementing BMPs for regulated stormwater are stormwater construction general permits, which increasingly require LID techniques (e.g., rain gardens, permeable pavement) for new development, and the MS4 permit, which sets stormwater limits for existing development. Management of
regulated urban stormwater is based on numeric, water quality-based pollutant wasteload allocations such as those published in the Chesapeake Bay TMDL; however, in the case of stormwater, permit limits may be expressed as implementation requirements in the form of BMPs.8 This is consistent with the Bay TMDL’s approach, and jurisdictional WIPs were also structured around BMP implementation to meet urban runoff allocations. However, stormwater rulemaking currently under development has proposed that numeric pollutant limits be incorporated into stormwater permits “where feasible.”9 If the final rule is adopted in accordance with this concept, stormwater general permits in TMDL implementation areas could be required to incorporate water quality-based effluent limits.
Watershed-based permitting can lead to cost savings as a consortium of permittees organize to distribute pollutant load allocations and contribute to monitoring and tracking efforts in their local or regional watersheds. An in-depth discussion of innovative stormwater management and regulatory permitting, including watershed-based permitting, is provided in Urban Stormwater Management in the United States (NRC, 2008).
Offsets for Growth and Development. Urban stormwater retrofits are much more costly and less efficient than BMPs developed as part of new development. Also, retrofits often rely on public funds, which can be difficult to procure and administer, while new development BMPs can be supported as part of the cost of construction. Cost savings from the use of LID techniques in new development compared to conventional BMP applications in existing development vary widely, but savings of 15 to 80 percent have been realized, with a few exceptions where LID technique costs exceeded conventional BMP costs (EPA, 2007). Therefore, a key goal for TMDL implementation should be to minimize increases in pollutant loads from new development to lessen the potential offset burden that would be required from existing development. One relatively easy urban/ suburban nutrient management strategy to implement is a “no net increase in nutrient loading” requirement, which would apply to new construction and redevelopment when various transactions occur, such as land sales, zoning changes, or land-use changes. To go a step further, communities might even require a percentage reduction in loadings associated with these occurrences, especially for redevelopment, as a means to attain load
8 November 22, 2002, memorandum from EPA, Robert H. Wayland, III and James A. Hanlon. Establishing Total Maximum Daily Load (TMDL) Wasteload Allocations (WLAs) for Storm Water Sources and NPDES Permit Requirements Based on Those WLAs.
9 November 12, 2010, EPA memorandum from James A. Hanlon and Denise Keehner, “Revisions to the November 22, 2002 Memorandum. “Establishing Total Maximum Daily Loads (TMDL) Wasteload Allocations (WLA) for Storm Water Sources and NPDES Permit Requirements Based on Those WLAs.”
reduction goals. Significant nutrient reductions can be realized when communities encourage redevelopment using LID techniques over new building on undeveloped lands.
As discussed earlier in this chapter, future population growth creates challenges for urban pollution control management. Under current strategies, existing WWTP discharge flows are lower than their permitted capacity (plant design capacity) to allow room for population growth. Urban strategies could be more aggressive if wastewater plants were required to offset any additional loads beyond current loads rather than permitted loads.
Residential Fertilizer Use. An estimated 9.5 percent (3.8 million acres) of the Bay watershed is in turf cover, with 75 percent of that in residential lawns. Assuming that 65 percent of turfgrass is fertilized with nitrogen lawn fertilizer, almost 215 million pounds of nitrogen per year could be applied within the watershed (Schueler, 2010). Fertilizer contribution within urban/suburban stormwater runoff typically ranges between 10 and 25 percent of total stormwater nutrient loads, depending on soil conditions, fertilizer application rate, and application timing prior to storm events (Barth, 1995; Linde and Watschke, 1997; Groffman et al., 2004; Shuman, 2004). Estimates from an urbanized watershed near Baltimore indicate that approximately 53 percent of the total nitrogen input was from the application of fertilizers (Groffman et al., 2004). Model-based estimates of nitrogen loading to the Chesapeake Bay from all sources indicate that fertilizer residues from developed landscapes account for about 10 percent of the loading (Figure 1-6). Boesch and Greer (2003) estimate that urban and suburban development in the Bay watershed could increase by more than 60 percent by 2030, and residential turfgrass coverage increases with urbanization. Therefore, implementing strategies to reduce the nutrient loads from residential fertilizer application could increase the likelihood of achieving the overall CBP nutrient reduction goals. Schueler (2010) states that “by changing attitudes and behaviors about what constitutes a green lawn, it may be possible to achieve major runoff and nutrient reductions.”
Restriction of residential fertilizer application and other landscape management actions have recently been enacted in New Jersey, Florida, Michigan, Minnesota, and Wisconsin (Box 5-5). Proponents of these efforts cite the cost-effectiveness per unit of nutrient loading reduced, community involvement, and educational opportunities as primary drivers. Although several states within the Bay Watershed have initiated some actions to address residential runoff through fertilizer management, most of the actions to date rely primarily on education, certification of lawn care professionals, and registration and labeling of residential fertilizer bags. New York is the
Examples of Residential Fertilizer Control Programs
New Jersey, Florida, Michigan, and Minnesota have recently enacted regulations to enhance water quality by restricting the application of residential fertilizer, and several have data showing the subsequent benefits of these regulations. In Minnesota, regional and state phosphorus fertilizer restriction laws were enacted in 2004 and 2005, respectively. 2006 estimates suggest that phosphorus fertilizer use (in tons) decreased by 48 percent after adoption of these laws. Much of these anticipated reductions in use were associated with the replacement and availability of phosphorus-free fertilizers at retail sales outlets (MDA, 2007).
A residential fertilizer law enacted in New Jersey in January 2011 requires that at least 20 percent of the nitrogen in all lawn fertilizers be in slow-release form, sets buffers between the turf on which fertilizer is applied and water bodies, prohibits the use of fertilizer during heavy rainfall, and bans the use of phosphorus in fertilizers (New Jersey Law A-2290).
To help manage nutrient loading to estuaries in Florida, several urbanized counties in southwest Florida have enacted residential fertilizer ordinances, which prohibit the application of phosphorus to lawns throughout the year (unless a soil test indicates the need for additional phosphorus) and nitrogen during the rainy summer months. Expected and observed water quality responses to residential fertilizer restrictions support consideration of their enactment. Modeled estimates of nitrogen loading under various residential fertilizer use scenarios indicate that moderate compliance (50 percent) with an ordinance that restricts the use of residential nitrogen fertilizer during the rainy season could reduce nitrogen loadings to Tampa Bay, Florida, by 4 percent of the urbanized load (TBEP, 2008).
Ann Arbor, Michigan, adopted an ordinance in 2006 that curtailed the use of phosphorus on lawns. After adoption, phosphorus levels in the Huron River dropped an average of 28 percent (Lehman et al., 2009).
exception: a 2010 law will prohibit the use of phosphorus-containing lawn fertilizer starting in 2012 unless establishing a new lawn or a test shows that the lawn is phosphorus-deficient. The law addresses phosphorus only because phosphorus is “usually the limiting nutrient in freshwater lakes” Chapter 205 of the Laws of New York, 2010ci). No prohibitions on nitrogen fertilizer application are currently in effect in the Bay jurisdictions although legislative proposals are currently being considered (A. Swanson, personal communication, Chesapeake Bay Commission, 2011).
Technologies are available for reliable nitrogen and phosphorus removal at wastewater treatment facilities at or below 3 mg/L and 0.1 mg/L, respectively (EPA, 2008b), and according to model estimates (see Appendix A) for 1985 to 2009, nitrogen and phosphorus loads in wastewater decreased by 42 and 59 percent, respectively. However, an additional 27 percent reduction for nitrogen and 26 percent reduction for phosphorus (relative to 2009 levels) will be required to meet the TMDL wasteload allocation for wastewater point sources. Assuming limits-of-technology can be implemented, combined sewer overflows (CSOs)10 abated, and performance sustained at wastewater plants throughout the watershed, the E3 model scenario (Appendix A) shows that nitrogen could be reduced by another 27 million lbs/yr and phosphorus by 3.27 million lbs/yr below the TMDL allocation. Costs of implementing these technologies remain the largest barrier.
Wastewater reuse for agricultural or urban landscape irrigation represents an untapped approach for further reducing wastewater nutrient loads. By beneficially reusing treated wastewater (also called reclaimed water) for irrigation, nutrient discharge is reduced and fertilizer applications on the irrigated lands can be reduced or eliminated (Asano et al., 2007). Landscape irrigation with reclaimed water is well accepted and widely practiced in the United States. Crook (2005) reported that more than 200 water reuse facilities provided reclaimed water for irrigation to more than 1,600 parks and playground. The treatment requirements for non-potable reuse for irrigation are not significantly greater than those already applied in most WWTPs in the Bay watershed, but the costs of distribution systems to supply reclaimed water to lands with large irrigation needs represent a significant barrier.
More than 2 million individual homeowner septic systems are estimated by the Watershed Model to contribute 4 percent of the nitrogen load to the Bay (Figure 1-6). Few septic systems have been upgraded to include nitrogen removal capability, a technology that can be costly to install and generally requires specialized operation and maintenance services. Most WIPs propose transitioning homes from septic systems to public wastewater collection and treatment systems as the most cost-effective means to reduce nitrogen loading from septic systems. However, public sewer infrastructure is often not accessible to more rural homeowners, and concerns that sewer systems will promote higher density growth, adding to pollutant loads from
10 Combined sewer systems collect stormwater and municipal sewage in the same infrastructure (i.e., pipes) for treatment at a wastewater treatment plant. Combined sewer overflows (CSOs) occur when heavy precipitation causes the total wastewater inflow to exceed the capacity of the treatment plant, and excess untreated wastewater is discharged directly into surface waters.
wastewater treatment plants have been expressed. Without septic system upgrades, some benefits in pollutant load reductions can be attained with proper maintenance practices, primarily inspection and regularly scheduled pumping, usually every 3 to 5 years. Some Bay jurisdictions require septic system permits or track maintenance within a regulatory framework to assure compliance. Upgraded systems for nitrogen removal are increasingly required, particularly when properties are in close proximity to receiving waters, when capacity for denitrification with groundwater transport is limited, or when new development occurs. In those cases, advanced treatment technologies may be beneficial and remove up to 50 percent of the nitrogen at the edge of the leach field if operated and maintained properly.
Enhancing Individual Responsibilities
To meet its nutrient and sediment reduction goals, the CBP must not only address large public or collective sources, such as sewage treatment plants, public lands and infrastructure, and agricultural and industrial entities, but everyday actions by watershed residents that are generally not regulated by law. Enhancing individual responsibilities, either through education, incentives, or regulations (e.g., restricting residential fertilizer use), can also contribute to the success of Bay restoration and to water quality improvements. Two areas already discussed where individual responsibilities could affect the Bay’s response include septic system maintenance and upgrades and residential landscape management. Two additional areas include residential improvements to reduce stormwater runoff and dietary changes.
Residential Actions to Reduce Stormwater. When used in residential landscape management, LID practices have demonstrated nutrient and sediment reduction capability that can achieve pre-development loading conditions in some cases. Because developed land is linked to excess nutrient and sediment loads (10, 31, and 19 percent of the total nitrogen, phosphorus, and sediment loads, respectively, according to model estimates; see Figures 1-6, 1-7, and 1-8), widespread application of LID practices on a voluntary basis or through incentive programs could benefit Bay water quality. Practices that promote infiltration (e.g., rain gardens, pervious alternatives to paving, rain barrels, natural landscaping techniques) reduce runoff and the nutrients and sediments associated with runoff (NRC, 2008). Although these practices may be implemented on a voluntary basis, incentive programs can promote more widespread recognition of the benefits of LID practices and their application. Similarly, LID techniques can become desirable lifestyle features (e.g., natural landscaping) and produce cost savings (e.g., rain water reuse, reduced fertilizer/pesticide use).
Dietary Changes. Human sewage, mostly from wastewater treatment plants, accounts for approximately 25 percent of the nitrogen to the Chesapeake Bay according to model estimates (Figure 1-6). The source of this nitrogen is protein consumed by watershed residents. Two diet-related actions would result in less nitrogen being injected into septic systems and municipal wastewater treatment plants. First, given that, on average, people in the United States consume more protein than is recommended in dietary reference intake guidelines (IOM, 2005), a decrease in protein consumption would decrease nitrogen discharges in wastewater. Second, a shift from to a less meat-intensive diet would reduce nitrogen losses to the environment during the food production process (Howarth et al., 2002).
Air Pollution Strategies
As noted in Figure 1-6, atmospheric sources are estimated to contribute 33 percent of the nitrogen loads to the Bay. The Chesapeake Bay has realized benefits from large decreases in NOx emissions from the Bay airshed resulting from the provisions of the Clean Air Act and its amendments, but the atmosphere is still a major source of nitrogen entering the Bay. Thus, more stringent controls on NOx emissions from all sources will benefit both the Bay and watershed residents, and benefits will exceed costs, primarily because healthcare costs attributed to air pollution will decrease (Birch et al., 2011). Examples of approaches to further reduce NOx emissions from power plants include operating installed NOx control equipment more frequently, using low sulfur coal, or installing additional control equipment (e.g., low NOx burners, selective catalytic reduction, or scrubbers) (EPA, 2010k).
Because of the significance of NHx deposition to the Bay and its watershed (estimated at more than 6 percent of the total nitrogen loads to the Bay;11 see Figure 1-6), controls on ammonia sources are also needed to reduce the significant impacts of crop and animal agriculture on the Bay. These controls are more difficult to implement for two reasons. First, there is no regulatory framework for ammonia, because it is not a criteria pollutant. Second, ammonia sources are diffuse, unlike the point sources of NOx. These two difficulties notwithstanding, efforts to decrease ammonia emissions could significantly reduce nitrogen deposition to the Bay and its watershed. Strategies for decreasing ammonia emissions focus on livestock dietary manipulations to lower the pH of manure and reduce the protein content in feed. Dietary changes can be made to shift a portion of the
11 The 6 percent of nitrogen from agricultural sources does not include NHx deposition that is part of the undifferentiated “atmospheric deposition to tidal waters,” which is estimated to make up 7 percent of the nitrogen loads to the Bay (Figure 1-6).
soluble nitrogen in urine to less soluble and more slowly degraded organic nitrogen in feces to decrease ammonia emissions (Powers, 2002). Additional strategies can be applied in manure management to reduce ammonia emissions, including applying chemical amendments that limit urea hydrolysis or lower the manure pH, minimizing moisture content, and using manure handling systems to separate feces from urine. Subsurface manure application using tillage equipment or injectors has been shown to reduce ammonia emissions compared to traditional surface application of manure. Additionally, covering manure storage areas can significantly reduce ammonia emissions, and ventilation systems with ammonia treatment systems can also be used (Powers, 2002; Becker and Graves, 2004).
As discussed previously in this chapter, the costs of meeting water quality goals pose considerable challenges to federal, state and local partners and to private individuals faced with changes to personal and business activities for implementation of water quality protection. How to pay for water quality protection, and other environmental public goods generally, is a perennial question. Although an extensive review of funding for CBP nutrient and sediment reduction strategies was beyond the committee’s charge, the committee did not want to raise the issue of high costs without also discussing some potentially viable funding strategies.
Targeting Agricultural BMP Funding
Many land-based best management practices are available that can result in nutrient and sediment load reductions. However, cost-share programs to encourage adoption of BMPs will be most cost-effective if they are targeted to locations where allocation of available funds results in the greatest load reductions possible. Nutrient and sediment load reductions and associated water quality improvement goals are most likely to be achieved if staff and monetary resources available to a given jurisdiction are targeted to a prioritized list of watersheds and their associated receiving waters. Priorities should reflect both the opportunities for nutrient and sediment reductions, such as hydrologically active areas of high nutrient or sediment source availability, and the costs of BMPs appropriate for such locations. Target locations may include agricultural areas with shorter flowpaths to reduce nitrogen losses and, for phosphorus, runoff-prone areas and areas with high phosphorus soils or excess manure relative to crop needs. Targeting strategies for reducing nutrient and sediment loads to the Bay can also consider other environmental and socioeconomic co-benefits, such as carbon sequestration and increased wildlife habitat and diversity in ripar-
ian buffers, although targeting strategies that attempt to address too many objectives may be weakened.
Evidence of successfully targeting financial assistance for conservation is not easy to find; payments to farmers have historically been distributed broadly with limited attention to potential for environmental benefit (Schertz and Doering, 1999; Lichtenberg and Smith-Ramirez, 2003). Not surprisingly, targeting limited federal and state resources can be controversial because access to cost-share assistance is not available equally to all farmers in all areas. However, Florida has successfully targeted state and USDA/NRCS funding to agricultural BMP cost-share programs in the Suwannee River and Lake Okeechobee Watersheds. Strong state/federal working relationships coupled with effective educational materials and outreach to landowners through the state Farm Bureau, commodity organizations, and land grant university-based cooperative extension service educators have proven effective. An adaptive management strategy can improve the effectiveness of a targeting program. If targeted monitoring reveals that nutrient and sediment load reductions are not achieved by control practices implemented in targeted watersheds, then evaluation of practice effectiveness and effectiveness of financial incentives for motivating practice adoption can be undertaken within the limited targeted area and the program adapted as appropriate (see also Chapter 4). Similar BMP performance evaluation and modification strategies are key elements of the agricultural nonpoint source control program employed in Florida, for example, described in Box 2-2.
Nutrient Offset and Credit Trading
With limited federal and state funding for financial assistance, the EPA, the U.S. Department of Agriculture (USDA), and states are relying on nutrient offset and credit trading programs as an alternative funding model to reduce point and nonpoint sources of nitrogen and phosphorus as required by the TMDL. However, views conflict about the extent to which trading can be relied upon to make a significant contribution to reducing loads and/ or to lowering the costs of reducing loads (Greenhalgh and Faeth, 2001; King and Kuch, 2003; Ribaudo et al., 2005; Shabman and Stephenson, 2007; Showalter and Spigener, 2007; Selman et al., 2010; Stephenson et al., 2010). Despite the development of almost 40 nutrient trading or offset programs across the United States, the number of successful completed trades is very small. In their nationwide review of programs, King and Kuch (2003) concluded that the few trades that have taken place have been primarily regulator-approved bilateral agreements negotiated between point source dischargers.
A number of supply and demand problems and institutional obstacles
stand to limit the success of nutrient trading programs. Selman et al. (2010) concluded that the demand for nutrient offsets and/or credits in the Bay region is likely to be strong, especially because of the expectation of growth in the watershed, but the supply of credits is more uncertain, largely because of requirements that both point sources and nonpoint sources meet baseline requirements before generating credits. A potential geographical mismatch between potential supply and demand may also be problematic. Selman et al. (2010) point to several examples of basins that exhibit the potential for either excess nutrient credit supply or excess demand that, in the absence of inter-basin trading, will not be captured in nutrient trading.
Neither supply nor demand is generally assured, however. The availability of nutrient credits is constrained by several factors, many of which reflect questions about whether CWA and TMDL language offer sufficient flexibility to make nutrient trading a viable option. Point sources may be reluctant to reduce effluent discharges below allowed levels to generate credits out of concern that implementing more aggressive controls signals that current technology-based or water quality-based effluent limits (which are supposed to reflect the maximum possible levels of control) could be adjusted downward to reflect the exhibited attainability of more stringent limits (Stephenson and Shabman, 2010). Nutrient credits from agricultural nonpoint sources may be fewer than expected because of requirements that the agricultural sources achieve minimum levels of nutrient reduction before credits are generated (King and Kuch, 2003).
Baseline requirements for nonpoint sources and analogous technology-based limits for point sources will raise the costs of generating credits. Point and nonpoint sources will use lower-cost reductions to meet baseline requirements, making additional reductions to generate credits more costly. Recent research in Virginia also questions whether sufficient credits can be produced by agricultural nonpoint sources when trading ratios (e.g., requiring reduction of 2 pounds of nonpoint source reduction to generate 1 pound of offset) and the sheer number of protected agricultural acres required to generate offsets are considered (Stephenson et al., 2010).
Demand side obstacles to nutrient credit trading may be more problematic, yet they are more subtle (King and Kuch, 2003). There is some evidence that point sources may look for lower-cost alternatives to purchasing nutrient credits, including water reuse, constructed wetlands, biomass harvest, and removal of on-site septic systems (Stephenson at al., 2010). Demand is further constrained when existing point sources are required to meet technology-based standards before purchasing nutrient credits. Requiring that nutrient trades become part of point sources’ NPDES permits may further dampen demand because risk-averse point sources may be reluctant to tie compliance with their NPDES permits to actions taken or not taken by a third party (King and Kuch, 2003; Stephenson et al., 2010).
Demand for nutrient credits to meet urban nonpoint-source limits may be limited because emerging stormwater programs are requiring developers to exhaust feasible on-site controls before purchasing nutrient credits (Stephenson and Shabman, 2010). New and expanding point sources are more likely than existing point sources to seek out nutrient credits to meet offset requirements (Selman et al., 2010; Stephenson et al., 2010). Even then, however, point sources may be hesitant to seek nutrient credits from nonpoint sources. Stephenson et al. (2010) determined, for example, that offsetting a discharge expansion of 1 million gallons per day (MGD)12 under Virginia’s nutrient trading program would require the application of continuous no-till on 10,000-25,000 acres of cropland. With average farm sizes ranging from 100 to 400 acres across Virginia’s four river basins, offsetting even a 1 MGD expansion would likely involve contracting with several dozen farm operations—a high transaction cost proposition for any individual point source.
Taken together, these challenges suggest that nutrient offset or credit trading is not a panacea for reaching nutrient reduction goals at lower cost. Removal of institutional constraints that restrict supply and demand at federal and state levels will be required if states are to implement effective trading programs (King and Kuch, 2003; Shabman and Stephenson, 2007).
Funding Urban Stormwater Management
Funding urban stormwater management is fraught with challenges but some innovative approaches are being considered. Increasingly, local entities are providing incentives to promote adoption of stormwater BMPs by homeowners and businesses, including LID techniques and other green infrastructure approaches (EPA, 2010j). Incentives are not always monetary; other forms of encouragement to promote BMP implementation include development incentives offered to developers during the process of applying for development permits, such as zoning upgrades, expedited permitting, and reduced regulatory requirements. Awards and recognition programs can also encourage homeowner and commercial efforts (EPA, 2010j). Supplementary granting programs, such as the federal Section 319 nonpoint source program, can help to defray implementation costs for unregulated stormwater activities, but at about $200 million/year nationally, available funds will not provide for all of the TMDL’s unregulated urban runoff control requirements.
12 For comparison, the Alexandria Sanitation Authority in Alexandria (Fairfax County), Virginia, processes on average 54 million gallons of wastewater per day and serves about 350,000 people in the City of Alexandria and part of Fairfax County (see http://www.alexsan.com/).
Section 319 funds cannot be used to meet requirements of the MS4 or other stormwater permits. To meet MS4 stormwater quantity and quality requirements, some municipalities have instituted stormwater or development fees that are assessed based on type of land use and area of impervious surface and increasingly administered through stormwater utilities (EPA, 2008c).
A stormwater utility (called a stormwater authority in Pennsylvania) is a mechanism to fund the cost of municipal services directly related to the control and treatment of stormwater. A stormwater utility will operate similarly as an electric or water utility. The utility will be administered and funded separately from the revenues in the general fund, ensuring a dedicated revenue source for the expense of stormwater management. (EPA, 2008c)
Generally, stormwater utilities collect fees from property owners based on the amount of stormwater runoff generated. Utilities commonly use an “equivalent residential unit” to establish fee rates, based on: (1) the amount of impervious cover in the parcel, regardless of size; (2) the intensity of development (i.e., the percentage of impervious cover relative to the entire parcel’s size); or (3) an equivalent hydraulic area, based on the combined impact of impervious and pervious cover within a parcel (EPA, 2008c).
Based on a Connecticut study, Fuss and O’Neill (2010b) concluded that stormwater utilities could effectively support implementation of LID by providing subsidies for LID demonstrations, funding for operation and maintenance, technical assistance in LID design and installation, and funding for retrofits for water quality improvements. The City of Portland, Oregon, instituted a Clean River Rewards Program that incentivizes participation by providing discounts on stormwater bills of homeowners who implement particular practices to “contain the rain” (City of Portland, 2006). Nationally, in 2009, stormwater utility fees varied widely, ranging from $8 to $160 per year for a single family home with an average fannual ee of $44 (Fuss and O’Neill, 2010b).
Funding Monitoring Strategies
As described in Chapter 4, monitoring and evaluation of reported ambient stream water quality, particularly at a small watershed scale, is a critical part of understanding the field-scale effectiveness and timescales of response following BMP implementation. Identifying sufficient funds to support an ambient monitoring program capable of detecting potential changes in local water quality in response to BMP implementation can be difficult for individual private entities, small municipalities, and even states.
NRCS has recently developed an interim Conservation Practice Standard (#799) to encourage monitoring and evaluation of BMP effectiveness by private landowners. It is being made available on a pilot basis, with 75 percent cost-sharing support through the NRCS Environmental Quality Incentives Program (EQIP), in a number of states that are part of the Mississippi River Basin Healthy Watershed Initiative (MRBI). The most landowner interest, to date, has been in Missouri, where state funds were used to cover the landowners’ portion of the costs and cover the technical expertise needed to implement monitoring protocols. However, overall, landowner participation has been limited (Thomas Christensen, NRCS, personal communication, 2011). No plans have been made to extend the pilot program to the Bay watershed states, although such an arrangement could be promising, particularly if coupled with targeted small-watershed monitoring initiatives that would complement the landowners’ edge-of-field monitoring. If applied in the Bay watershed, collaboration between landowners and state or federal agency representatives or university scientists would be needed to develop monitoring plans and to install equipment.
One example of a successful local government collaboration to provide financial support for ambient water quality monitoring is the Southern California Stormwater Monitoring Coalition (SMC). The SMC was formed in 2001 as part of the Southern California Coastal Water Research Project (SCCWRP), a collaborative public agency created in 1969 to conduct coastal environmental monitoring and research.13 The SMC was the result of a cooperative agreement among the Phase I municipal stormwater NPDES lead permittees, the NPDES regulatory agencies in Southern California, and SCCWRP.14 The SMC members agreed, with EPA cooperation, that NPDES compliance monitoring schedules would be adjusted periodically to make available funding that may be appropriately re-directed to support cooperative ambient monitoring and reporting efforts.
To enhance support for CBP monitoring efforts (and adaptive management), local and regional governments and industries within Bay subwatersheds may wish to consider similar cooperative efforts. One option is re-directing some funds currently used for individual NPDES compliance monitoring toward an established localized ambient monitoring program or a collaborative effort between different monitoring programs using standardized data collection approaches to allow data collation and comparison. Similarly, some percentage of federal funds provided to agricultural and other landowners to cost-share BMP implementation could be directed to existing or collaborative ambient water quality monitoring programs specifically designed to detect potential changes in stream water quality
associated with BMP implementation. As noted in Chapter 4, such monitoring programs would need to be carefully targeted toward addressing specific uncertainties related to practice effectiveness and Bay response if the monitoring is to support adaptive management.
Establishing a Chesapeake Bay Modeling Laboratory
The final strategy that the committee presents in this chapter addresses improving the scientific and modeling support for the CBP to increase the likelihood that the program will meet its ultimate goal—recovery of the Chesapeake Bay. The committee was not asked to—and did not—review the models. However, the models that collectively make up the Chesapeake Bay Model (i.e., the Airshed Model, the Watershed Model, and the Bay Model; see Box 1-1) are central to the proper allocation of restoration resources, evaluation and planning, and the ongoing adaptive management of the Bay in a changing future. The models have been used to estimate the loading reductions of nitrogen, phosphorus, and sediment necessary to achieve water quality and living resources objectives (i.e., the TMDLs). Models are used to estimate the effect of BMPs on loading reductions to the Bay, thereby providing essential information for planning and evaluating implementation strategies. Models are central to forecasting the Bay’s response to future loading reductions and to system perturbations, such as climate change and annual differences in precipitation to the watershed. Thus, models are essential to the success of the CBP. As a consequence, they need to be continuously evaluated as new data are collected, updated as mechanistic understanding increases, and scrutinized for inconsistencies and possible computational and scientific inaccuracies.
The models presently reside in two locations: the Watershed Model at the EPA Chesapeake Bay Program Office and the Bay Model at the U.S. Army Waterways Experiment Station. Only a few technical professionals are completely familiar with the details of the models, their history of development, and the long series of changes and improvements that have been made over the 25 years of development. This is a fragile and precarious situation. Although the codes for both models are publicly available at the Chesapeake Community Modeling Program (CCMP), they are complex and using them would pose a challenge for even experienced modelers. There is no active community involved in exercising these models. The documentation that exists for the models is no substitute for a community of scientists and engineers who understand their inner workings and have actually used the models. This is in sharp contrast to other modeling communities, for example the climate modeling community, in which there are multiple modeling efforts, some of which are centered in national laboratories, and for which comparisons of the various models is a common practice (e.g.,
Macadam et al., 2010). The Chesapeake Bay hydrodynamic modeling community shares some of these characteristics, but the Airshed, Watershed, and Bay Models do not.
The CCMP has held and continues to hold open and regular meetings during which progress in building, calibrating, and using the models is discussed. However, these models are not used by academics for research investigating the mechanisms that control ecological responses of the Bay. The number of persons at present who can actually make computations is limited to a very few, and there are just two senior scientists among the CBP modeling group—one each for the Watershed Model and Bay Model. Their time in the past has been completely committed (actually over-committed) to the tasks associated with building, calibrating, and using the models to fulfill various management requirements, most recently the development of the TMDL. As a consequence time has been unavailable for the critical cooperative work with the scientific community that would enable a much wider familiarity with and acceptance of these models.
Credibility of these models among the scientific, engineering, and management communities that are concerned with understanding, managing, and protecting Bay water quality is critically important. A recent analysis by LimnoTech (2010), which used a USDA model designed to simulate changes in nutrient loading resulting from conservation practices on crop land in the Bay watershed, reported discrepancies with the CBP Watershed Model, including the amount of agricultural nutrients that reach the Bay. The LimnoTech report has fueled a growing backlash against the Bay TMDL and spurred several members of the House Agriculture Committee to conclude that the CBP models used to develop the TMDL are “fatally flawed” (Blankenship, 2011). Although this NRC committee did not analyze the LimnoTech report or the discrepancies in the models, this issue highlights how technical concerns regarding the CBP models can undermine support for the CBP goals and strategies, the details of which are developed and evaluated using the models. Because the models are not widely used outside the CBP, they lack credibility with the broader scientific community that would result from a history of independent applications. Thus, the academic community has largely been unable to weigh in on this recent controversy, although the CBP Scientific and Technical Advisory Committee is planning an independent review of the LimnoTech report.
Considering the magnitude of the remediation costs and the value of the Bay resource, this situation needs to be addressed. The atmospheric and oceanographic communities have national laboratories (i.e., the National Center for Atmospheric Research [NCAR] and the Geophysical Fluid Dynamics Laboratory [GFDL], respectively) that are centers for the development of atmospheric and oceanic circulation models and more recently
the climate models that are used to forecast the possible consequences of various climate-related control measures. A similar laboratory entrusted with the stewardship of the Chesapeake Bay models could be developed for the CBP and charged with evaluating monitoring data and uncertainty in model simulations, improving the predictive skill of the models, and continuously seeking model improvements to accommodate new scientific understanding of the system. Such a laboratory could also be central in designing and improving the CBP monitoring programs, evaluating the consequences of adaptive management experiments, helping to understand where and why pollution controls did not perform as effectively as planned, identifying science gaps, and evaluating the consequences of climate change. Finally, it would be the place where sound technical analysis and advice could be obtained by managers for the inevitable changes that will be necessary as nutrient and sediment reductions are implemented and the resulting responses of the Bay ecosystem are evaluated.
When specific issues are raised, smaller scale models built to answer specific questions could be implemented and/or developed as part of the laboratory’s research. A lab would have the personnel to do the development and would not be wed to one watershed model or one Bay model. A laboratory could also facilitate improvements to the models to support the 2017 re-evaluation of the TMDL and the WIPs.
Involvement of the academic community in a laboratory is vitally important. The flow of ideas among the policy, management, and academic communities is a crucial part of the continuing development of state-of-the-art models and understanding. Faculty could form research associations with lab personnel, and lab personnel could have appointments in academic departments. The success of NCAR and GFDL is due in part to their proximity to research universities. Recognition of the need for improved integration of the academic community and the CBP modeling program is not new.15 What is new is the recommendation that an actual laboratory be established that fulfills the functions listed above and is more than just a virtual association of collaborating individuals. Instead, the committee envisions a modeling laboratory as a physical location, following the examples of NCAR and GFDL.
The actual institutional sponsorship of the laboratory, its relationship to management agencies, and the makeup of the research staff would require serious deliberation. There are tradeoffs to be considered. A lab that is too “academic” might not be responsive to immediate needs. A focus that is too “operational” would merely continue the current situation where scientific functions are not given sufficient priority. A lab with too many varied responsibilities would dilute the effort from a focus on modeling.
Surveying similar labs and their successes and failures would be a useful exercise. The NOAA Great Lakes Environmental Research Laboratory, the EPA research labs, and the Everglades Interagency Modeling Center are additional examples worth examining.
An important component of the work of a modeling laboratory would be the integration of monitoring with modeling efforts, as recommended in Chapter 4. A laboratory could contribute to designing future data collection efforts, relocating sampling stations where the uncertainty of the Bay response is largest, locating monitoring stations in the watershed where loading reductions are predicted to have the largest observable changes, and supporting adaptive management experiments. Because monitoring is costly, any improvement in existing monitoring efficiency could make resources available for other needs.
Integrated modeling and monitoring is also needed to help determine whether CBP management actions are working as anticipated (STAC, 1997, 2005), and this requires models that can accurately simulate the time scales of BMP response and nutrient storage and transport. Time lags between land-based BMP implementation in the Bay watershed and full responses in nutrient and sediment loadings (see Box 1-3), however, remain poorly understood and have not been quantified. The existing models incorporate some of the necessary mechanisms, but others are clearly missing or are not well calibrated. For example, the Bay Model includes a sediment model that is capable of calculating lag times associated with the degradation of organic nitrogen and the storage of inorganic phosphorus (see Figure 1-2), but the land simulation in the Watershed Model has no routing from the land surface to the streams to account for nutrient storage in soils, nor a sediment model for the stream beds, and thus no associated lag times. BMPs are, instead, modeled as instantaneous and permanent pollutant reductions. Also, the groundwater lag in the Watershed Model is virtually nonexistent (hours to days). To incorporate this time lag would require coupling to a separate groundwater model to simulate lags based on groundwater flow (G. Shenk, CBPO, personal communication, 2011). Increases in nitrogen in the Choptank River, as described by Hirsch et al. (2010; see Figure 1-12b) are representative of groundwater and surface water interactions that are not simulated well by the Watershed Model.
Through a Chesapeake Bay modeling laboratory, disciplinary scientists and engineers and modelers could collaborate to quantify lag times in the Bay watershed and translate the phenomena into operational calculation frameworks. Additional intensive monitoring in small watersheds could be conducted to quantify the time scales of contributing mechanisms. Model hindcasting could be used to analyze whether existing models are capable of accurately forecasting the course of Bay remediation and elucidate the
strengths and weaknesses of the present formulations.16 If deemed necessary, additional smaller-scale models could be developed to simulate the time frames of BMP responses. This research would be essential to respond to concerns that management plans are not performing as expected and to support the analysis of progress. Additionally, if significant lag times between implementation of land-based BMPs and nutrient loads reductions are determined, the research could help maintain public support for continued efforts and investments in Bay recovery.
Reaching the long-term CBP nutrient and sediment reduction goals will require substantial commitment from each of the Bay jurisdictions and likely some level of sacrifice from all who live and work in the watershed. Jurisdictions not only need to significantly reduce current loads, but they will need to take additional actions to address future growth and development over the next 15 years. Additionally, the Bay partners will need to adapt to future changes (e.g., climate change, changing agricultural practices) that may further impact water quality and ecosystem responses to planned implementation strategies. To reach the long-term load reduction goals, Bay jurisdictions and the federal government will need to prepare for the challenges ahead and consider a wide range of possible strategies, including some that are receiving little, if any, consideration today.
Success in meeting CBP goals will require careful attention to the consequences of future population levels, development patterns, agricultural production systems, and changing climate dynamics in the Bay Watershed. Nutrient and sediment management efforts are taking place in the context of a quickly changing landscape and uncertain outcomes that could significantly affect the strategies needed to attain the TMDL goals. For example, an increase in the concentration of livestock or dairy animals near processing and distribution centers would mean a greater concentration of manure nutrients in these areas than has existed in the past. Additionally, Bay jurisdictions may need to adjust future milestone efforts to larger than anticipated population and more intensive land-use development scenarios, as well as climate change influences. Further and continued study of future scenarios is warranted to help Bay partners adapt to a changing future.
Helping the public understand lag times and uncertainties associated with water quality improvements and developing program strategies to
16 For example, starting with 1950 simulations, the CBP could calculate the relationship between loadings to the bay and the water quality responses from 1950 to the present. The computations can then be verified against observations to better understand the lag times incorporated in the model.
account for them are vital to sustaining public support for the program, especially if near-term Bay response does not meet expectations. Although the science and policy communities generally recognize the uncertainties inherent in water quality modeling, load projections, and practice effectiveness and expect that water quality successes will lag implementation, the same may not be true of the broader public. If the public expects visible, tangible evidence of local and Bay water quality improvements in fairly short order, they will almost certainly become frustrated. In the absence of a concerted effort to engage Bay residents in a conversation about the dynamics of the Bay and how and when improvements can be expected, CBP partners should anticipate and be prepared to respond to an impatient or disillusioned public. By developing small watershed-scale monitoring efforts that highlight local-scale improvements and associated time lags in water quality as they occur, the CBP can better understand and inform the public about anticipated responses to, and expectations for, nutrient control measures.
The committee identified potential strategies that could be used by the CBP partners to help meet their long-term goals for nutrient and sediment reduction and ultimately Bay recovery. The committee did not attempt to identify every possible strategy that could be implemented but instead focused on approaches that are not being implemented to their full potential or that may have substantial, unrealized potential in the Bay watershed. Because many of these strategies have policy or societal implications that could not be fully evaluated by the committee, the strategies are not prioritized but are offered to encourage further consideration and exploration among the CBP partners and stakeholders. Examples include:
• Improved and innovative manure management. Possible strategies include expanded CAFO permitting programs, guidelines and/or regulations to control the timing and rates of manure application, innovative manure application methods, transport of manure to watersheds with the nutrient carrying capacity to accept it, alternative uses (e.g., bioenergy production), animal nutrition management to reduce nutrient loading, and limits on the extent of animal operations based on the nutrient carrying capacity of the watershed.
• Incentive-based approaches and alternative regulatory models. Several approaches have been used successfully elsewhere to increase the use of agricultural BMPs for the purpose of improving water quality. Florida developed a voluntary, incentive-based BMP program that provides regulatory relief in exchange for BMP implementation, maintenance, and reporting. Denmark’s nutrient management program provides an alterna-
tive model that couples agricultural regulatory requirements with incentives and has resulted in large reductions in nutrient surpluses. The CBP could facilitate an analysis of the costs and potential effectiveness of various incentive-based and regulatory alternatives.
• Regulatory models that address stormwater, growth and development, and residential fertilizer use. Watershed-based permitting for urban stormwater can lead to cost savings if a consortium of permittees chooses to organize to distribute pollutant load allocations and contribute to monitoring and tracking efforts in their local or regional watersheds. Restrictions on nitrogen and phosphorus residential fertilizer application are cost-effective methods of nutrient load management in urban and suburban areas. Communities could also adopt regulations to restrict land-use changes that would increase nutrient loads from stormwater runoff or cap wastewater treatment plant discharges at current levels, requiring offsets for any future increases.
• Enhanced individual responsibility. Enhancing individual responsibilities, either through education and incentives or through regulations, can also contribute to the success of Bay restoration and to water quality improvements. Examples of actions that individuals can take to improve water quality include increasing application of low-impact design and residential stormwater controls, changing residential landscape management, maintaining and upgrading septic systems, and changing diets.
• Additional air pollution controls. Although the Chesapeake Bay has realized substantial benefits from the Clean Air Act, the atmosphere remains a major source of nitrogen entering the Bay. More stringent controls on nitrogen emissions from all sources, including NOx and agricultural ammonia emissions, will benefit both the Bay and the people who reside in its watershed.
Innovative funding models will be needed to address the expected costs of meeting Bay water quality goals. Targeting agricultural BMP cost-share programs is not always politically popular, but it can produce greater reductions at lower cost than will distributing resources broadly with little attention to water quality impacts. Although nutrient trading among point and nonpoint sources is often cited as a mechanism to reach nutrient reduction goals at lower cost, its potential for reducing costs is limited. Stormwater utilities offer a viable funding mechanism to support stormwater management efforts of municipalities. Funding for monitoring will also be needed,
and successful regional monitoring cooperatives in other parts of the United States may be useful models.
Establishing a Chesapeake Bay modeling laboratory would ensure that the CBP would have access to a suite of models that are astate-of-the-art and could be used to build credibility with the scientific, engineering, and management communities. The CBP relies heavily on models for setting goals and evaluating nutrient control strategies; thus, the models are essential management tools that merit substantial investment to ensure that they can fulfill present and future needs. Currently, only a few technical professionals are fully knowledgeable of the details of the models and their development. The models are not widely used outside the CBP and, therefore, are unfamiliar to the broader scientific community. Credibility of the models is essential if the CBP goals and strategies are to be accepted and have widespread support. A Chesapeake Bay modeling laboratory would bring together academic scientists and engineers with CBP modelers to examine various competing models with similar objectives and work to enhance the quality of the simulations. An important component of the work of a modeling laboratory would be the integration of monitoring with modeling efforts. Joint research investigations focused on evaluating the success of the Bay recovery strategies could be centered in the laboratory, such as studies on the role of lag times in the observed pollutant loads and Bay responses. A close association with a research university would bring both critical review and new ideas. A laboratory could also facilitate improvements to the models to support the 2017 re-evaluation of the TMDL and the WIPs.