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Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution (2000)

Chapter: 5 Sources of Nutrient Inputs to Estuaries and Coastal Waters

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Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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5
Sources of Nutrient Inputs to Estuaries and Coastal Waters

KEY POINTS IN CHAPTER 5

This chapter reviews the sources and amounts of nutrients supplied to coastal water bodies and finds:

  • Globally, human activity has dramatically increased the flux of phosphorus (by a factor of almost 3) to the world’s oceans. There has been an even more dramatic increase in nitrogen flux, especially in the last 40 years, with the greatest flux adjacent to areas of highest population density. Human activity has increased the flux of nitrogen in the Mississippi River by some 4-fold, in the rivers in the northeastern United States by some 8-fold, and in the rivers draining to the North Sea by more than 10-fold.

  • Although point source nutrients are the major problem for small watersheds adjacent to major population centers, these inputs are relatively easy to minimize with tertiary wastewater treatment processes. In contrast, nutrients from nonpoint sources have become the dominant and least easily controlled component of nutrients transported into coastal waters from large watersheds, and especially from watersheds with extensive agricultural activity or atmospheric nitrogen pollution.

  • Phosphorus flux to estuaries is dominantly derived from agricultural activities as particle-bound forms mobilized in runoff. In some areas, groundwater transported phosphorus is also important.

  • Nitrogen input to estuaries is derived from both agricultural activity (e.g., dominant in the Mississippi River) and fossil-fuel combustion (e.g., dominant in the northeastern United States). Animal feeding operations have become a major contributor to nitrogen exports.

  • It is likely that the atmospheric component of nitrogen flux into estuaries has previously been under-estimated. This component is derived from fossil-fuel

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

combustion and from animal feedlots and other agricultural sources, and is both deposited directly into estuaries and also deposited initially onto the land surface and then carried into estuaries by runoff.

Human activity has an enormous influence on the global cycling of nutrients, especially on the movement of nutrients to estuaries and other coastal waters. For phosphorus, global fluxes are dominated by the essentially one-way flow of phosphorus carried in eroded materials and wastewater from the land to the oceans, where it is ultimately buried in ocean sediments (Hedley and Sharpley 1998). The size of this flux is currently estimated at 22 Tg P yr−1 (Howarth et al. 1995). Prior to increased human agricultural and industrial activity, the flow is estimated to have been around 8 Tg P yr−1 (Howarth et al. 1995). Thus, current human activities cause an extra 14 Tg of phosphorus to flow into the ocean sediment sink each year, or approximately the same as the amount of phosphorus fertilizer (16 Tg P) applied to agricultural land each year.

The effect of human activity on the global cycling of nitrogen is equally immense, and furthermore, the rate of change in the pattern of use is much greater (Galloway et al. 1995). The single largest global change in the nitrogen cycle comes from increased reliance on synthetic inorganic fertilizers, which accounts for more than half of the human alteration of the nitrogen cycle (Vitousek et al. 1997). The process for making inorganic nitrogen fertilizer was invented during World War I, but was not widely used until the 1950s. The rate of use increased steadily until the late 1980s, when the collapse of the former Soviet Union led to great disruptions in agriculture and fertilizer use in Russia and much of eastern Europe. These disruptions resulted in a slight decline in global nitrogen fertilizer use for a few years (Matson et al. 1997). By 1995, the global use of inorganic nitrogen fertilizer was again growing rapidly, with much of the growth driven by increased use in China (Figure 5-1). Use as of 1996 was approximately 83 Tg N yr−1. Approximately half of the inorganic nitrogen fertilizer that was ever used on Earth has been applied during the last 15 years.

Production of nitrogen fertilizer is the largest process whereby human activity mobilizes nitrogen globally (Box 5-1). However, other human-controlled processes, such as combustion of fossil fuels and production of nitrogen-fixing crops in agriculture, convert atmospheric nitrogen into biologically available forms of nitrogen. Overall, human fixation of nitrogen (including production of fertilizer, combustion of fossil fuel, and production of nitrogen-fixing agricultural crops) increased globally some two- to three-fold between 1960 to 1990 and continues to grow (Galloway et al. 1995). By the mid 1990s, human activities made new nitrogen avail-

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

FIGURE 5-1 Annual global nitrogen fertilizer consumption for 1960-1995 (1 Tg = 1012 g; data from FAO 1999). The rate of increase was relatively steady until the late 1980s, when collapse of the former Soviet Union reduced fertilizer use in Russia. Fertilizer use is growing again, driven in large part by use in China (modified from Matson et al. 1997).

BOX 5-1
The Fate of Nitrogen Fertilizer in North America

When nitrogen fertilizer is applied to a field, it can move through a variety of flow paths to downstream aquatic ecosystems (Figure 5-2). Some of the fertilizer leaches directly to groundwater and surface waters, with the range varying from 3 percent to 80 percent of the fertilizer applied, depending upon soil characteristics, climate, and crop type (Howarth et al. 1996). On average for North America, some 20 percent is leached directly to surface waters (NRC 1993a; Howarth et al. 1996). Some fertilizer is volatilized directly to the atmosphere; in the United States, this averages 2 percent of the application, but the value is higher in tropical countries and also in countries that use more ammonium-based fertilizers, such as China (Bouwman et al. 1997). Much of the nitrogen from fertilizer is incorporated into crops and is removed from the field in the crops when they are harvested, which is of course the objective of the farmer. A recent National Research Council report

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

(NRC 1993a) suggests that on average 65 percent of the nitrogen applied to croplands in the United States is harvested, although other estimates are somewhat lower (Howarth et al. 1996). By difference, on average approximately 13 percent of the nitrogen applied must be building up in soils or denitrified to nitrogen gas.

Since much of the nitrogen is harvested in crops, it is important to trace its eventual fate. The majority of the nitrogen is fed to animals (an amount equivalent to 45 percent of the amount of fertilizer originally applied, if 65 percent of the nitrogen is actually harvested in crops; Bouwman and Booij 1998). Some of the

FIGURE 5-2 The average fate of nitrogen fertilizer applied to agricultural fields for North America. The numbers in parentheses are calculated by difference, and the other numbers are direct estimates (unpublished figure by R. Howarth).

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

nitrogen is directly consumed by humans eating vegetable crops—in North America perhaps 10 percent of the amount of nitrogen originally applied to the fields (Bouwman and Booij 1998). By difference, perhaps 10 percent of the amount of nitrogen originally applied to fields is lost during food processing, being placed in landfills or released to surface waters from food-processing plants.

Of the nitrogen that is consumed by animals, much is volatilized from animal wastes to the atmosphere as ammonia. In North America, this volatilization is roughly one-third of the nitrogen fed to animals (Bouwman et al. 1997), or 15 percent of the amount of nitrogen originally placed on the fields. This ammonia is deposited back onto the landscape, often near the source of volatilization, although some of it first travels for long distances through the atmosphere (Holland et al. 1999). Some of the nitrogen in animals is consumed by humans, an amount roughly equivalent to 10 percent of the amount of nitrogen fed to the animals, or 4 percent of the nitrogen originally applied to fields. By difference, the remainder of the nitrogen—over 25 percent of the amount of nitrogen originally applied to the fields—is contained in animals wastes that are building up somewhere in the environment. Most of this may be leached to surface waters.

Of the nitrogen consumed by humans, either through vegetable crops or meat, some is released through wastewater treatment plants and from septic tanks. In North America, this is an amount equivalent to approximately 5 percent of the amount of nitrogen originally applied to fields (Howarth et al. 1996). By difference, the rest of the nitrogen is placed as food wastes in landfills or is denitrified to nitrogen in wastewater treatment plants and septic tanks.

The conclusion is that fertilizer leaching from fields is only a portion of the nitrogen that potentially reaches estuaries and coastal waters. Probably of greater importance for North America as a whole is the nitrogen that is volatilized to the atmosphere or released to surface waters from animal wastes and landfills. Since food is often shipped over long distances in the United States, the environmental effect of the nitrogen can occur well away from the original site of fertilizer application.

able at a rate of some 140 Tg N yr-1 (Vitousek et al. 1997), matching the natural rate of biological nitrogen fixation on all the land surfaces of the world (Vitousek et al. 1997; Cleveland et al. 1999). Thus, the rate at which humans have altered nitrogen availability globally far exceeds the rate at which humans have altered the global carbon cycle (Figure 5-3).

The human alteration of nutrient cycles is not uniform over the earth, and the greatest changes are concentrated in the areas of greatest popula-

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

FIGURE 5-3 The relative change in nitrogen fixation caused by human activities globally compared to the increase in carbon dioxide in the atmosphere since 1900. Note that humans are having a much greater influence on nitrogen availability than they are on the production of carbon dioxide, an important greenhouse gas (modified from Vitousek et al. 1997).

tion density and greatest agricultural production. Some regions of the world have seen very little change in the flux of either nitrogen or phosphorus to the coast (Howarth et al. 1995, 1996), while in other places the change has been tremendous. Human activity is estimated to have increased nitrogen inputs to the coastal waters of the northeastern United States generally, and to Chesapeake Bay specifically, by some six- to eight-fold (Boynton et al. 1995; Howarth et al. 1996; Howarth 1998). Atmospheric deposition of nitrogen has increased even more than this in the northeast (Holland et al. 1999). The time trends in human perturbation of

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

nutrient cycles can also vary among regions. For example, while the global use of inorganic nitrogen fertilizer continues to increase, the use of nitrogen fertilizer in the United States has increased relatively little since 1985 (Figure 5-4; Evans et al. 1996).

Note, however, that the use of nitrogen fertilizer in the United States in the next century may again increase to support greater exports of food to developing countries. Countries such as China have been largely self sufficient in food production for the past two decades, in part because of increased use of nitrogen fertilizer. The use of fertilizer in China is now very high—almost 10-fold greater than in the United States—and further increases in fertilizer use are less likely to lead to huge increases in food production as they have in the past. Therefore, if China’s population continues to grow it may once again be forced to import food from the United States and other developed countries, leading to more use of nitrogen fertilizer here.

WASTEWATER AND NONPOINT SOURCE INPUTS

Traditionally, most water quality management emphasizes control of discharges from wastewater treatment plants and other point sources.

FIGURE 5-4 U.S. commercial fertilizer use (modified from Evans et al. 1996).

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

However, generally of greater concern for nutrients and coastal eutrophication are “nonpoint sources” of nutrients (NRC 1993a). A regional-scale analysis of fluxes of nitrogen from the landscape to the coast of the North Atlantic Ocean demonstrated that nonpoint sources of nitrogen exceeded sewage inputs for all regions in both Europe and North America (Howarth et al. 1996). Overall, sewage contributed only 12 percent of the flux of nitrogen from the North American landscape to the North Atlantic Ocean (Howarth et al. 1996). Nonpoint sources also dominate for phosphorus inputs to surface waters in the United States (Sharpley and Rekolainen 1997; Carpenter et al. 1998), and because of an effort to control phosphorus point source pollution, nonpoint sources of phosphorus have grown in relative importance since 1980 (Jaworski 1990; Sharpley et al. 1994; Litke 1999).

Wastewater inputs can sometimes be a major source of nitrogen to an estuary when the watershed is heavily populated and small relative to the surface area of the estuary itself (Nixon and Pilson 1983). Even in some estuaries fed by larger watersheds, wastewater can be the largest source of nitrogen if the watershed is heavily populated. For example, wastewater contributes an estimated 60 percent of the nitrogen inputs to Long Island Sound, largely due to sewage from New York City (CDEP and NYSDEC 1998). However, nitrogen and phosphorus inputs from nonpoint sources in most estuaries are greater than are inputs from wastewater, particularly in estuaries that have relatively large watersheds (NRC 1993a). For example, only one-quarter of the nitrogen and phosphorus inputs to Chesapeake Bay come from wastewater treatment plants and other such point sources (Boynton et al. 1995; Nixon et al. 1996). For the Mississippi River, sewage and industrial point sources contribute an estimated 10 percent (Howarth et al. 1996) to 20 percent (Goolsby et al. 1999) of the total nitrogen flux (organic and inorganic nitrogen) and 40 percent of the total phosphorus flux (Goolsby et al. 1999).

As discussed in more detail in Chapter 9, many technologies exist for reducing nutrient discharges from wastewater treatment plants. The relatively standard approaches of using primary and secondary sewage treatment lower phosphorus and nitrogen discharges on average by approximately 20 percent to 25 percent, although there is a significant variation among plants (Viessman and Hammer 1998; NRC 1993a). Additional tertiary treatment for nutrient removal can lower nitrogen discharges by 80 percent to 88 percent and phosphorus discharges by 95 percent to 99 percent (NRC 1993a). However, most wastewater treatment plants in the United States do not have adequate nitrogen removal capabilities. In Tampa Bay, wastewater treatment plants were a major source of nitrogen prior to the institution of tertiary nitrogen removal, and this treatment has

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

successfully reversed the trend in eutrophication there (Johansson and Greening 2000).

Reduction in the eutrophication of most estuaries requires the management of nutrient inputs from nonpoint sources in addition to those of wastewater treatment plants and industrial sources (NRC 1993a). The nature of these sources is described in the remainder of this chapter.

DISTURBANCE, NONPOINT NUTRIENT FLUXES, AND BASELINES FOR NUTRIENT EXPORTS FROM PRISTINE SYSTEMS

In a landscape that is completely undisturbed by humans, export of nitrogen and phosphorus to downstream aquatic ecosystems tends to be small, particularly in the temperate zone (Hobbie and Likens 1973; Omernik 1977; Rast and Lee 1978; Howarth et al. 1996). Assuming that the landscape is in an ecological steady state, the export of nutrients cannot exceed the inputs. For nitrogen, these inputs are biological nitrogen fixation and deposition of nitrogen compounds from the atmosphere; in the temperate zone both tend to be small in the absence of human disturbance (Howarth et al. 1996; Cleveland et al. 1999; Holland et al. 1999). Thus, the export of nitrogen from undisturbed temperate landscapes must also be low, in fact lower than the input because there is some accumulation of nitrogen in the system and some loss of nitrogen through denitrification (the bacterial conversion of reactive nitrate into nonreactive molecular nitrogen). For tropical regions, rates of biological nitrogen fixation and natural deposition of nitrogen from the atmosphere are far higher, and so nitrogen export to downstream ecosystems from undisturbed ecosystems may also be greater (Howarth et al. 1996; Cleveland et al. 1999; Holland et al. 1999; Lewis et al. 1999).

Unfortunately, it is difficult to determine with any precision the magnitude of the natural flux of nitrogen from a temperate landscape like the United States. Atmospheric pollution and the resulting elevated nitrogen deposition are widespread, providing some level of disturbance virtually everywhere in the country, and in fact in most of the world’s temperate ecosystems (Holland et al. 1999). There are a few remaining temperate forests that do not receive elevated nitrogen deposition from pollution sources, such as some remote forests in Chile (Hedin et al. 1995). However, these are poor models for most of the temperate systems of the United States as the Chilean forests receive high precipitation and runoff, and have vastly different ecological histories.

An expert panel under the auspices of the International SCOPE (Scientific Committee on Problems of the Environment) Nitrogen Project estimated that pristine temperate-zone ecosystems, such as those that had

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

characterized much of North America and Europe prior to human disturbance, would export between 75 and 230 kg N km−2 yr−1 to downstream aquatic ecosystems, with the median estimate being 133 kg N km−2 yr1 (Howarth et al. 1996; Howarth 1998). This provides the best estimate available for the natural, background load of nitrogen from the landscapes in the continental United States. Valigura et al. (2000) estimate that for estuaries with small watersheds, the nitrogen flux off the pristine landscape prior to European colonization was at the low end of this range, perhaps 78 to 108 kg N km−2 yr−1. Assuming a baseline flux of 133 kg N km−2 yr−1 for an undisturbed temperate landscape, human activity has increased the nitrogen flux in the Mississippi River by more than 4-fold, in the rivers of the northeastern United States by 8-fold, and in the rivers draining to the North Sea by 11-fold (Howarth 1998). In an independent analysis for Chesapeake Bay, Boynton et al. (1995) estimated that nitrogen fluxes have increased some 6- to 8-fold since pre-colonial times, a value consistent with the conclusion from the International SCOPE Nitrogen Project.

In an undisturbed landscape, the major source of phosphorus to a terrestrial ecosystem is the weathering of the soil and parent-rock material, which tends to be relatively slow and therefore sets a low limit on the export of phosphorus. As a global average, the export of phosphorus from the terrestrial landscape prior to human disturbance can be estimated from the oceanic sedimentary record and was somewhat greater than 50 kg P km−2 yr−1, expressed per area of land surface (Howarth et al. 1995). However, this clearly depends on the phosphorus content of the parent-rock material, the rate of weathering, and other environmental conditions, including the rate of erosion. The current flux of phosphorus from the landscape is in fact less than 50 kg P km−2 yr−1 for more than half of the area in the Mississippi River basin (Goolsby et al. 1999), and is only 5 kg P km−2 yr−1 for the watersheds of Hudson’s Bay, Canada (Howarth et al. 1996). On the other hand, the rather large export of phosphorus from the Amazon River basin of over 230 kg P km−2 yr−1 appears to be a largely natural phenomenon (Howarth et al. 1996). Given the site-specific nature of phosphorus export and the paucity of information on background phosphorus losses from a given location prior to cultivation, no baseline for the natural rate of phosphorus export exists.

Disturbance of the landscape increases the export of both nitrogen and phosphorus, although there are some major differences in the responses of these two nutrients. As a general rule, most export of phosphorus from disturbed systems occurs as phosphorus bound to particles, so factors regulating erosion and sedimentation are critical in controlling phosphorus fluxes. An important exception can occur in sandy soils with

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

low phosphorus adsorption capacities (Sharpley et al. 1998; Sims et al. 1998); such soils can be important to consider when managing eutrophication in portions of the Atlantic coastal plain. Some phosphorus moves through the atmosphere as dust particles, and this can contribute greatly to the phosphorus economies of some remote oceanic waters and forests. In general, such inputs of phosphorus to estuaries and coastal waters are not as important as inputs in surface waters.

For nitrogen, some export also occurs in particle-bound forms, but nitrogen tends to be much more mobile through soils in dissolved form than phosphorus, so significant exports can occur in groundwater (Paerl 1997) or as dissolved nitrogen in surface waters. In addition (and also unlike phosphorus), reactive nitrogen compounds can be quite mobile in the atmosphere. For example, significant amounts of ammonia gas from agricultural sources (particularly urea- and ammonia-based fertilizers, manures, and animal feedlot wastes) volatilize to the atmosphere and are deposited elsewhere in the landscape (Bouwman et al. 1997; Holland et al. 1999). Globally, of the 60 to 80 Tg N yr−1 applied as inorganic nitrogen fertilizer, 21 to 52 Tg N yr−1 are estimated to be volatilized to the atmosphere as ammonia, either directly from the fertilizer or from animal waste (Holland et al. 1999). That is, on average some 40 percent of the inorganic nitrogen fertilizer that is applied cycles through the atmosphere and is redeposited. In the United States, the value is somewhat lower, but still 25 percent of the inorganic nitrogen fertilizer that is used is volatilized to the atmosphere (Holland et al. 1999).

For phosphorus, agriculture is the largest disturbance controlling nonpoint fluxes of phosphorus in the landscape (Carpenter et al. 1998). For nitrogen, both agriculture and fossil-fuel combustion contribute significantly to nonpoint source flows to estuaries and coastal waters (Howarth et al. 1996). Some of this nitrogen export comes directly from agricultural fields, but because of both substantial nitrogen transport in the atmosphere and nitrogen mobility in dissolved forms, the nitrogen export from other types of ecosystems, including forests, can be substantial. Since agriculture dominates the nonpoint source flux of phosphorus and contributes significantly to nonpoint sources of nitrogen (often dominating it as well), changes in agricultural practices over the last few decades contribute to these nutrient fluxes. Industrial and fossil fuel sources of nitrogen and the mechanisms that control both nitrogen and phosphorus fluxes in the landscape will be discussed later in this chapter.

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

CHANGES IN AGRICULTURAL PRODUCTION AND NONPOINT SOURCE NUTRIENT POLLUTION

One of the greatest changes in agriculture has been the use of inorganic fertilizers, which expanded dramatically after World War II in response to the demand for increased agricultural output. In the developed countries, large processing plants were built to manufacture nitrogenous fertilizers and convert imported rock phosphate into a variety of water-soluble and partially water-soluble phosphorus fertilizer products. Basic slag, a by-product from the steel industry, also became widely used in the manufacturing of phosphorus fertilizer. In the United States, the use of inorganic phosphorus fertilizer rose rapidly in the 1940s and 1950s, but has been relatively constant since 1960. The rate of use of inorganic nitrogen fertilizer, on the other hand, continued to rise rapidly until the early 1980s (Figure 5-4). This relative gain in nitrogen use over phosphorus use resulted primarily from favorable crop yield responses, especially corn, to nitrogen fertilizers.

Over the last 30 years, agricultural production systems in the United States have become more specialized and concentrated. During this time, overall agricultural production has more than doubled (Evans et al. 1996), and is occurring on less agricultural land and on fewer but larger farms (Evans et al. 1996). Since 1950, U.S. farmland has decreased from 1,200 to 970 million acres (20 percent) and the number of farms has dropped from 5.6 to 2.1 million (63 percent), while average farm size has increased from 213 to 469 acres (120 percent).

In many states, animal feeding operations (AFOs) are now a major source of agricultural income. The rapid growth of the animal industry in certain areas of the United States has been coupled with an intensification of operations. For example, current census information shows an 18 percent increase in the numbers of hogs in the United States over the last 10 years, at the same time as a 72 percent decrease in numbers of hog farms. Over the same 10 years, the number of dairy farms decreased 40 percent, but herd size increased 50 percent. A similar intensification of the poultry and beef industries has also occurred, with 97 percent of poultry production in the United States coming from operations with more than 100,000 birds and over a third of beef production coming from just under 2 percent of the feedlots (Gardner 1998). Driving this intensification is an increased demand for animal products and improved profitability because of advances in transportation, processing, and marketing. But animal feeding operations pose significant challenges with the management of wastes produced.

Prior to World War II, farming communities tended to be self-sufficient, in that enough feed was produced locally to meet animal requirements

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

and the manure nutrients could be effectively recycled to meet crop needs. After World War II, increased fertilizer use in crop production fragmented farming systems and created specialized crop and animal operations that efficiently coexisted in different regions. Since farmers did not need to rely on manures as fertilizers (the primary source until fertilizer production and distribution became cheap), grain and animal production could be spatially separated. By 1995 the major animal producing states imported over 80 percent of their grain for feed (Lanyon and Thompson 1996). In fact, less than a third of the grain produced on farms today is fed to animals on the farm where it is grown (USDA 1989).

This evolution of agricultural systems is resulting in a major transfer of nutrients from grain-producing areas to animal-producing areas and, consequently, accumulation of nitrogen and phosphorus in soils of the animal-producing areas. For example, the potential for nitrogen and phosphorus surplus at the farm scale can be much greater in AFOs than in cropping systems, because nutrient inputs become dominated by feed rather than fertilizer (Isermann 1990; NRC 1993b). Thus, many water quality concerns are a result of this imbalance in system inputs and outputs of nitrogen and phosphorus, which have been brought about by an increase in AFOs. Lander et al. (1998) calculated the amounts of nitrogen and phosphorus produced by manure in confined AFOs on a countywide basis (Figure 5-5). From this and crop yield information, Lander et al. (1998) were able to identify those counties where more than 100 percent of the nitrogen and phosphorus needed for crop production was available from livestock manure. (In other words, counties where manure production exceeded crop need—assuming that manure was applied only to non-legumes and harvested crop land and hay land; Figure 5-6).

The number of U.S. counties where manure nitrogen and phosphorus exceeds the potential crop uptake and removal has been steadily increasing since 1950 (Figure 5-7). This increase has been greater for phosphorus than nitrogen (Kellogg and Lander 1999). In those areas with an excess of nitrogen and phosphorus relative to crop needs, there is a greater risk of nutrient export from agricultural watersheds to surface and ground waters (Figure 5-8). This excess of nutrients in manure tends to occur in areas where downstream export is likely due to relatively wet climates, since high water availability is conducive to animal feeding operations.

The limited large-scale geographic information available to summarize phosphorus soil test results shows trends in soil phosphorus buildup to very high levels in some areas. These areas of phosphorus build up and often coincide with areas of intensive animal production (Fixen 1998; Figure 5-9). Soils in this category require little or no input of phosphorus, either from fertilizer or organic by-products, for economically optimum crop production. In many of these areas, this build-up of soil phosphorus

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

FIGURE 5-5 Estimated manure nitrogen and phosphorus production from confined livestock (modified from Lander et al. 1998).

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

FIGURE 5-6 Potential for nitrogen and phosphorus available from animal manure to meet or exceed plant uptake and removal on non-legume, harvested cropland, and hayland (modified from Lander at al. 1998).

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

FIGURE 5-7 Number of counties where manure nutrients exceed the potential plant uptake and removal, including pastureland application (modified from Kellogg and Lander 1999).

has increased the risk for phosphorus movement in surface runoff and in some cases (notably Florida and the Delmarva Peninsula) into shallow groundwater aquifers. Unfortunately, many of these areas of soil phosphorus build-up tend to occur in areas with phosphorus-sensitive water resources or major drainage ways, such as the inland waters of the Carolinas, Florida Everglades, Lake Okeechobee, Great Lakes, and the Mississippi River basin (Figure 5-9). Although this survey of soil phosphorus includes only samples sent for analysis and does not represent a complete survey of all soils in the United States, it does highlight some of the effects of long-term changes on agricultural production systems.

How has this come about? Using the Chesapeake Bay drainage basin as an example, if all the manure produced within the basin in 1939 were made available for application to corn, large areas of the corn cropland would not have received adequate amounts (3,000 kg km−2 yr−1; Figure 5-10; Lanyon 1999). Some areas of the basin with higher potential applications, such as New York and western Virginia, probably had limited areas of corn production at that time. In other words, without importation of fertilizer from outside the basin, the availability of manure limited crop

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

FIGURE 5-8 Watersheds with a high potential for soil and water degradation from manure nitrogen and phosphorus (modified from Kellogg and Lander 1999).

production. However, by 1992 manure production exceeded the corn requirements in large areas of Virginia, West Virginia, Delaware, parts of Pennsylvania, and the New York area of the drainage basin (Figure 5-10), and the need to dispose of excess manure (rather than crop need) began to shape patterns of fertilizer application. These patterns of nutrient distri-

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

FIGURE 5-9 Percent of soil samples analyzed in state laboratories that tested in the “high or above” range for phosphorus in 1997. Highlighted states had more than 50 percent of soil samples testing in the “high or above” range (modified from Fixen 1998).

bution and accumulation have come about by a number of complex and interrelated factors, not merely independent farmer decisions (Lanyon 1999). Farmers could do little to increase nutrient supplies on their farms when nitrogen and phosphorus fertilizers were scarce. It was only after the emergence of the fertilizer industry and the associated pattern of intensive animal feed operations that nitrogen and phosphorus supplies on farms could be increased to exceed farm nutrient requirements.

Export of Phosphorus from Agricultural Systems

Several surveys of U.S. watersheds have clearly shown that phosphorus loss in runoff increases as the forested portion of the watershed decreases and agriculture increases (Omernik 1977; Rast and Lee 1978). In general, forested watersheds conserve phosphorus, with phosphorus input in dust and in rainfall usually exceeding outputs in stream flow (Taylor et al. 1971; Hobbie and Likens 1973; Schreiber et al. 1976). Surface runoff from forests, grasslands, and other noncultivated soils carries little sediment,

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

FIGURE 5-10 Available manure phosphorus per acre of corn in the Chesapeake Bay drainage basin before and after World War II (1939 and 1992, respectively) (modified from Lanyon 1999).

so phosphorus fluxes are low and the export that occurs is generally dominated by dissolved phosphorus. This loss of phosphorus from forested land tends to be similar to that found in subsurface or dissolved base flow from agricultural land (Ryden et al. 1973; House and Casey 1988). The cultivation of land in agriculture greatly increases erosion and with it the export of particle-bound phosphorus. Typically, particulate fluxes constitute 60 to 90 percent of phosphorus exported from most cultivated land (Sharpley et al. 1995). In the eastern United States, conversion of land from forests to agriculture between 1700 and 1900 resulted in a 10-fold increase in soil erosion and a presumed similar increase in phosphorus export to coastal waters, even without any addition of phosphorus fertilizer (Meade 1988; Howarth et al. 1996). The soil-bound phosphorus includes both inorganic phosphorus associated with soil par-

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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ticles and phosphorus bound in organic material eroded during flow events. Some of the sediment-bound phosphorus is not readily available (Howarth et al. 1995), but much of it can be a long-term source of phosphorus for aquatic biota (Sharpley 1993; Ekholm 1994).

Increases in phosphorus export from agricultural landscapes have been measured after the application of phosphorus (Sharpley and Rekolainen 1997). Phosphorus export is influenced by the rate, time, and method of phosphorus application, form of fertilizer or manure applied, amount and time of rainfall after application, and land cover. These losses are often small from the standpoint of farmers (generally less than 200 kg P km−2), and represent a minor proportion of fertilizer or manure phosphorus applied (generally less than 5 percent). Thus, these losses are not of economic importance to farmers in terms of irreplaceable fertility. However, they can contribute to eutrophication of downstream aquatic ecosystems.

While phosphorus export from agricultural systems is usually dominated by surface runoff, important exceptions occur in sandy, acid organic, or peaty soils that have low phosphorus adsorption capacities and in soils where the preferential flow of water can occur rapidly through macropores (Sharpley et al. 1998; Sims et al. 1998). Soils that allow substantial subsurface export of dissolved phosphorus are common on parts of the Atlantic coastal plain and in Florida, and are thus important to consider in the management of coastal eutrophication in these regions.

Although there exists a good understanding of the chemistry of phosphorus in soil—water systems, the hydrologic pathways linking spatially variable phosphorus sources, sinks, temporary storages, and transport processes in landscapes are less well understood. This information is critical to the development of effective management programs that address the reduction of phosphorus export from agricultural watersheds.

Runoff production in many watersheds in humid climates is controlled by the variable source area concept of watershed hydrology (Ward 1984). Here, surface runoff is usually generated only from limited source areas in a watershed. These source areas vary over time, expanding and contracting rapidly during a storm as a function of precipitation, temperature, soils, topography, ground water, and moisture status over the watershed (Gburek and Sharpley 1998). Surface runoff from these areas is limited by soil-water storage rather than infiltration capacity. This situation usually results from high water tables or soil moisture contents in near-stream areas.

The boundaries of surface runoff-producing areas will be dynamic both in and between rainfalls (Gburek and Sharpley 1998). During a rainfall, area boundaries will migrate upslope as rainwater input increases. In dry summer months, the runoff-producing area will be closer to the

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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stream than during wetter winter months, when the boundaries expand away from the stream channel.

Soil structure, geologic strata, and topography influence the location and movement of variable source areas of surface runoff in a watershed. Fragipans or other layers, such as clay pans of distinct permeability changes, can determine when and where perched water tables occur. Shale or sandstone strata also influence soil moisture content and location of saturated zones. For example, water will perch on less permeable layers in the subsurface profile and become evident as surface flow or springs at specific locations in a watershed. Converging topography in vertical or horizontal planes, slope breaks, and hill slope depressions or spurs, also influence variable source area hydrology in watersheds. Net precipitation (precipitation minus evapotranspiration) governs watershed discharge and thus total phosphorus loads to surface waters. This should be taken into account when comparing the load estimates from different regions. It is also one reason why there seems to be more concern with phosphorus in humid regions than in more arid regions.

In watersheds where surface runoff is limited by infiltration rate rather than soil-water storage capacity, areas of the watershed can alternate between sources and sinks of surface flow. This again will be a function of soil properties, rainfall intensity and duration, and antecedent moisture condition. As surface runoff is the main mechanism by which phosphorus is exported from most watersheds (Sharpley and Syers 1979), it is clear that, if surface runoff does not occur, phosphorus export can be small.

Export of Nitrogen from Agricultural Systems

The fate of nitrogen applied as fertilizer to agricultural fields has received extensive study. Overall, nitrogen use in agriculture tends to be relatively inefficient (less than 25 percent of that applied), with animal uptake particularly small (less than 20 percent) compared with crop production systems (NRC 1993b). Generally for the United States, 45 percent to 75 percent of the nitrogen in fertilizer is removed in crop harvest (Bock 1984; Nelson 1985; NRC 1993b). Of the remainder, some is stored as organic nitrogen in the soil, some is volatilized to the atmosphere, and some leaches to ground and surface waters. A variety of factors, including soil type, climate, fertilizer type, and farming practices, influence the fate of fertilizer use (Howarth et al. 1996). For typical farming practices in the United States, the percentage of fertilizer that leaches to ground and surface waters varies between 10 and 40 percent for loam and clay soils, and 25 and 80 percent for sandy soils (Howarth et al. 1996). Overall in North America, it is estimated that 20 percent of the fertilizer nitrogen

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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applied to agricultural fields leaches into ground and surface waters (Howarth et al. 1996), although much of that is lost to denitrification in downstream wetlands, streams, and rivers before reaching estuaries or coastal waters.

A variety of factors affect the volatilization of nitrogen from fertilizer to the atmosphere, including soil type, climate, farming practices, and type of fertilizer (Bouwman et al. 1997). For example, when ammonium sulfate is applied to a soil with a pH below 5.5, less than 2 percent of the ammonium is volatilized (in the form of ammonia) to the atmosphere. Conversely, when ammonium sulfate is applied to calcareous soil (which has a higher pH), up to 50 percent of the nitrogen can be volatilized as ammonia gas to the atmosphere (Whitehead and Raistrick 1990; Bouwman et al. 1997). For typical farming practices, climate, and soils in the United States and Europe, Bouwman et al. (1997) estimated that on average 8 percent of the nitrogen in ammonium sulfate and 15 percent of the nitrogen in urea is volatilized to the atmosphere. The percentages are greater in tropical countries, and the volatilization from nitrate-based fertilizers is much less. While emissions of nitric oxide to the atmosphere are an important nitrogen loss from fertilized fields in tropical areas, this is generally a very small flux in temperate regions, including the United States (Holland et al. 1999). Virtually all the nitrogen volatilized from agricultural fields is eventually redeposited back onto the landscape and can reach estuaries and coastal waters (Howarth et al. 1996). Generally, this nitrogen is redeposited quite close to the point of emission (Holland et al. 1999).

Since 45 to 75 percent of the nitrogen applied as fertilizer is harvested in crops, tracing the fate of nitrogen in food and feedstock is important for understanding nitrogen inputs to natural waters (Howarth et al. 1996). The nitrogen in foods that are consumed by humans becomes sewage and is released in sewage effluent, where it is volatilized to the atmosphere as ammonia from sewage treatment plants or is denitrified (converted to plant-unavailable nitrogen) in the sewage treatment plants. However, in the United States most crops are fed to animals (Bouwman and Booij 1998). Thus, most of the nitrogen in harvested crops is excreted by animals. For animals such as poultry, hogs, and cows kept in barns or sheds, 36 percent of the excreted nitrogen on average is volatilized to the atmosphere as ammonia; keeping cows in meadows instead of barns reduces the atmospheric volatilization by more than 50 percent (Bouwman et al. 1997).

Assuming that (1) 65 percent of the nitrogen applied as fertilizer is removed in crops (NRC 1993b); (2) two-thirds of the crop production in the United States is fed to animals (Bouwman and Booij 1998); (3) the nitrogen growth efficiency for animals is 10 percent (Bouwman and Booij

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

1998); and (4) 36 percent of the nitrogen excreted by animals is volatilized to the atmosphere (Bouwman et al. 1997), then some 14 percent of all nitrogen applied in fertilizer is eventually volatilized to the atmosphere as ammonia after being consumed by animals. This is in addition to direct volatilization of ammonia from fertilizers and from sewage treatment plants.

Stated another way, ammonia volatilization to the atmosphere from agricultural systems in the United States is of the same order of magnitude as nitrate leaching from agricultural fields into surface waters. In addition, although losses are poorly documented, animal wastes also contribute nitrogen directly to surface waters (Howarth 1998). In a regional comparison of nitrogen cycling in major regions of the United States and Europe, Howarth (1998) found that estimates of nitrogen consumption by domestic animals were far better as predictors of nonpoint source nitrogen fluxes in rivers than were rates of application of inorganic nitrogen fertilizer.

Fate of Nitrogen in Atmospheric Deposition

Reactive nitrogen in the atmosphere includes both reduced compounds (NHy) and oxidized compounds (NOy). These come from a variety of sources, including fossil-fuel combustion, biomass burning, lightning, and emissions from soils. In the United States, most NOy comes from fossil-fuel combustion and most NHy comes from emissions from agricultural sources (Howarth et al. 1996; Prospero et al. 1996; Bouwman et al. 1997; Holland et al. 1999). The lifetime in the atmosphere for many of these reactive nitrogen compounds is short—from hours to a few days—and a large portion of the nitrogen is deposited near its source (Holland et al. 1999). NOy contributes to “acid rain,” but estuarine waters are well buffered and are not directly susceptible to acidification. Thus, the threat from NOy discussed here is its role as a contributor of nitrogen for coastal eutrophication.

Nitrogen deposition directly onto the water surfaces of estuaries and coastal waters can be substantial, although this is difficult to measure. Monitoring stations for atmospheric input of nitrogen tend to be scarce in coastal areas (Chapter 8). Where monitoring stations exist, they tend to measure only the nitrogen deposited in precipitation (wet deposition). Dry deposition of nitrogen (the impaction of particles and gases of nitrogen onto water, plant, or land surfaces) has proven difficult to measure in any type of ecosystem, and usually only wet deposition or at best some portion of dry deposition are measured at monitoring sites (Holland et al. 1999).

Evidence indicates that deposition directly onto the water surfaces of

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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estuaries tends to contribute from 1 percent to 40 percent of the total nitrogen inputs (Nixon et al. 1996; Paerl 1997; Paerl and Whitall 1999; Valigura et al. 2000), with estuaries such as the Baltic Sea (Nixon et al. 1996) and Tampa Bay (Zarbock et al. 1996) at the upper end of this range. Furthermore, evidence suggests a significant movement of nitrogen in the atmosphere from the eastern United States to the coastal and even offshore waters of the North Atlantic Ocean where it is deposited (Prospero et al. 1996; Holland et al. 1999); this flux could be as large as half the entire amount of reactive nitrogen emitted into the Earth’s atmosphere from the United States. However, because of the large natural flux of nitrogen from the deepwater of the North Altantic Ocean onto the continental shelf off the eastern U.S., this atmospheric deposition probably contributes less than 10 percent of the total input of nitrogen to the surface waters of the continental shelf (Howarth 1998).

Much of the reactive nitrogen deposited from the atmosphere falls onto terrestrial ecosystems. This can affect estuaries and coastal waters to the extent that it is exported from land. The fate of nitrogen deposition in forests has received extensive study. Productivity of most U.S. forests in their natural state is limited by the supply of nitrogen (Vitousek and Howarth 1991). As more nitrogen is made available to these forests from atmospheric deposition, production and storage of nitrogen in organic matter can be expected to increase temporarily. However, this ability of forests to store nitrogen is limited. Forests become “saturated” with respect to nitrogen when inputs exceed the total amount needed by trees and the assimilation capacity (through microbial and abiotic processes) of soil organic matter (Aber et al. 1989; Gundersen and Bashkin 1994; Magill et al. 1997; Emmett et al. 1998). Once a forest is saturated with respect to nitrogen, losses both to the atmosphere and to downstream ecosystems can increase rapidly. In European forests that have received high levels of nitrogen deposition for some time, the downstream export of nitrogen can be high, often greater than 500 kg N km−2 yr−1 (van Breement et al. 1982; Hauhs et al. 1989; Schulze et al. 1989; Durka et al. 1994). Some evidence indicates that the process whereby forests switch from retaining nitrogen to exporting nitrogen as they become nitrogen saturated can be self-accelerating due to related changes in biogeochemical cycling and ecosystem decline (Schulze et al. 1989; Howarth et al. 1996).

Ecological theory suggests that young aggrading forests tend to retain more nitrogen and be less likely to become nitrogen saturated than old-growth mature forests (Vitousek and Reiners 1975; Aber et al. 1989). Therefore, forests that have been logged or burned within the past several decades to a century can be expected to retain more nitrogen from deposition. However, a variety of factors, in addition to land-use history, can affect the ability of a forest to retain nitrogen, including the species com-

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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position of trees, climate, and soil type (Howarth et al. 1996; Aber and Driscoll 1997; Aber et al. 1997; Magill et al. 1997; Emmett et al. 1998). For example, Lajtha et al. (1995) found that only about half the nitrogen input from atmospheric deposition was retained by forests at Cape Cod, Massachusetts, whether the forests were young or mature, apparently because the sandy soils there allow nitrogen to pass through so quickly. On the other hand, forests on stony and sandy loam soils in western Massachusetts retained over 85 percent of nitrogen inputs even when heavily fertilized with nitrogen over a six-year period (Magill et al. 1997). In this fertilization experiment, there was some evidence that nitrogen retention decreased over time as the nitrogen content of the forest increased and that nitrogen saturation occurs more rapidly in pine forests than in hardwood forests (Magill et al. 1997).

In a review of nitrogen retention in U.S. forests, Johnson (1992) found no relationship between nitrogen inputs and nitrogen losses to downstream ecosystems—the percentage of nitrogen deposition that was retained varied among forests from nearly none to virtually all. Much of this variation could have been caused by differences in land use, soil type, and dominant tree species (Lajtha et al. 1995; Aber and Driscoll 1997; Magill et al. 1997; Emmett et al. 1998). Some of the variation, however, could have been due to the exclusion of dissolved organic nitrogen fluxes in the budgets considered by Johnson (1992), all of which included losses only of inorganic nitrogen; losses of organic nitrogen can be considerable from some forests (Hedin et al. 1995; Lewis et al. 1999). In fact, even in northern New England where atmospheric deposition of nitrogen is moderately high, most of the dissolved nitrogen leaving forests is organic nitrogen, rather than inorganic nitrogen (Campbell et al. 2000). Further, many of the budgets reviewed by Johnson (1992) were based on short-term studies, and losses of nitrogen from forests can show considerable year-to-year variation in response to climatic variation (Aber and Driscoll 1997). Finally, dry deposition of nitrogen is difficult to estimate precisely (Howarth et al. 1996; Holland et al. 1999; Valigura et al. 2000), and may not have been correctly characterized in some of the budgets summarized by Johnson (1992).

Emmett et al. (1998) proposed that the extent of nitrogen leaching from a forest can be easily predicted from the “nitrogen status” of the forest, as measured by the ratio of organic carbon to nitrogen in the forest floor. They experimentally illustrated that forests with a low nitrogen status (forest floor carbon:nitrogen greater than 30:1) retain most of the nitrogen added (well over 90 percent), whereas forests with a high nitrogen status (forest floor carbon:nitrogen less than 25:1) retain less than half the nitrogen added through deposition and fertilizer. Similarly, Campbell et al. (2000) have demonstrated that the ratio of organic carbon to organic

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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nitrogen in streamwater draining forests in the northeastern United States is a good predictor of export of inorganic nitrogen from those forests. For example, export of inorganic nitrogen increased dramatically in systems where the organic carbon to organic nitrogen ratio of the streamwater was below 20:1 to 25:1 (Campbell et al. 2000). A variety of factors control whether a forest acts as a sink or source of nitrogen, but forests experiencing long periods of high nitrogen inputs through atmospheric deposition will tend to become saturated with respect to nitrogen (Emmett et al. 1998). Thus, over time a forested watershed that experiences high inputs of nitrogen deposition will reach its capacity to store nitrogen and will begin to act as a source of nitrogen to the streams that drain it. Thus, its not surprising that experiments showed that nitrogen leaching from a forest slowed quickly after deposition was reduced through the use of roof exclosures (Bredemeier et al. 1998). This lead Emmett et al. (1998) to suggest that “immediate benefits in water quality could be expected following any reduction in nitrogen deposition loading.”

For management of lake acidification in some areas of Europe, managers have adopted the “critical load” concept (Bashkin 1997). This approach sets a goal of keeping atmospheric deposition below some level where it is thought that downstream release will be kept small enough to keep any ecological damage at an acceptable level. Research supports the conclusion that downstream release of nitrogen (and associated acid) can be expected to occur when a critical load value of 1,000 kg N km−2 yr1 from atmospheric deposition is reached (Schulze et al. 1989; Pardo and Driscoll 1993; Emmett and Reynolds 1996; Williams et al. 1996; Skeffington 1999; Figure 5-11). Average levels of nitrogen deposition (wet plus dry) currently exceed 1,000 kg N km−2 yr−1 for the northeastern United States and for much of Europe (Howarth et al. 1996; Prospero et al. 1996; Holland et al. 1999; Valigura et al. 2000).

The export of nitrogen following deposition onto terrestrial ecosystems other than forests has received less study. Some evidence indicates that grasslands are as retentive of nitrogen as forests, or even more so (Dodds et al. 1996). When nitrogen from the atmosphere is deposited onto agricultural fields, its fate is similar to the fate of nitrogen fertilizer applied to such fields, although nitrogen deposited during the non-growing season could be prone to greater loss to downstream ecosystems. The fate of atmospheric nitrogen deposited onto urban and suburban landscapes appears to be virtually unstudied, although the nitrogen content of stormwater runoff from urban environments is high (EPA 1983). Deposition onto urban landscapes is high, as expected since much of the reactive nitrogen in the atmosphere is deposited near sources, and it is reasonable to expect that the export of this deposition to coastal waters is also high. However, nitrogen deposition (wet or dry) in urban environments is

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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FIGURE 5-11 Ambient inputs (throughfall) and leaching losses at the Nitrogen Saturation Experiment sites (modified from Emmett et al. 1998).

poorly measured since most deposition monitoring sites are in rural environments (Holland et al. 1999). Uncertainty over the extent of nitrogen deposition in urban environments is one of the greatest uncertainties in the nitrogen budget for the United States (Holland et al. 1999).

PROCESSING OF NITROGEN AND PHOSPHORUS IN WETLANDS, STREAMS, AND RIVERS

Not all the nitrogen or phosphorus that is exported from a forest, a corn field, or an animal feedlot will reach an estuary or coastal waters because significant processing of nutrient flows can occur in wetlands, streams, lakes, reservoirs, and rivers that lie between the terrestrial systems and coastal waters (Kirchner and Dillon 1975; Kelly et al. 1987; Howarth et al. 1995, 1996; Rigler and Peters 1995). The sediments in wetlands and aquatic systems sometimes retain phosphorus and sometimes they do

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

not, depending on the presence of such phosphorus-sorptive phases as iron (III) hydroxides and oxides and calcium carbonate minerals (Howarth et al. 1995). Where iron compounds are the dominant phosphorus-adsorbing minerals, the oxidation state of the minerals is important, as oxidized iron compounds sorb phosphorus and reduced iron compounds do not (Theis and McCabe 1978; Howarth et al. 1995). Biological controls on the pH of the immediate surface sediment can also be important (Knuuttila et al. 1994). Despite these details, the retention of phosphorus in lakes can often be predicted from the residence time of water in the lakes, and lakes often retain 80 percent or more of the entering phosphorus (Kirchner and Dillon 1975; Rigler and Peters 1995; Nixon et al. 1996). Wetlands also frequently retain significant quantities of phosphorus, so buffer or riparian zones around streams or waterbodies can reduce inputs from agricultural land (Lowrance et al. 1984a, b, 1985).

Wetlands and aquatic systems can also remove significant quantities of nitrogen, although the mechanism is different. Generally, most removal of nitrogen in these systems is by denitrification. The process occurs largely in anoxic environments, and often occurs at high rates in the sediments of wetlands, lakes, and rivers. Most studies show a large removal of nitrate in groundwater that flows through wetlands, with most of this presumed to be denitrified (Peterjohn and Correll 1984; Lowrance et al. 1984b; Correll et al. 1992; Jordan et al. 1993; Jansson et al. 1994; Vought et al. 1994; Howarth et al. 1996), although some studies show a conversion to organic nitrogen (Devito et al. 1989; Brunet et al. 1994). Riparian wetlands that intercept waters flowing from agricultural fields before they enter streams can be effective at lowering nitrogen loads. However, in many areas these riparian wetlands were drained or otherwise destroyed as land was converted for agriculture (Vought et al. 1994). Krug (1993) has shown that in southern Sweden, the conversion of the last 10 percent to 15 percent of land into agricultural use disproportionately destroyed fringing wetlands and therefore doubled nitrogen inputs to streams. Wetland restoration has been suggested as both the cheapest and most effective approach for lowering nitrogen fluxes through the landscape to rivers (Rosenberg et al. 1990; Haycock et al. 1993).

Denitrification occurs in the water column of streams and rivers when the water is anoxic or extremely hypoxic. However, with the general improvement in river water quality that accompanied the widespread use of secondary sewage treatment for removal of biological oxygen demand, few rivers in the United States now have such low-oxygen events (NRC 1993a). On the other hand, the sediments of streams and rivers are frequently anoxic even when water quality is high, and this provides an ideal location for denitrification. Similarly, the sediments of lakes are almost always anoxic below the first few meters of water. For lakes,

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

streams, and rivers with an oxic water column, the extent of removal of nitrogen by denitrification can be modeled as a function of depth and mean residence time (Kelly et al. 1987; Howarth et al. 1996); denitrification is greater in shallower systems and in systems with a longer water residence time, since both of these lead to greater contact of nitrogen in the water with anoxic sediments. Using this model, Howarth et al. (1996) suggested that denitrification in river systems in the United States are in general unlikely to denitrify more than 20 percent of the nitrogen that flows into them.

NUTRIENT FLUXES TO THE COAST

Insights from a Regional Analysis

At the scale of individual estuaries, it has proven exceedingly difficult to determine the ultimate source for nitrogen inputs and the magnitude of the load each source contributes. Numerous obstacles exist to understanding nutrient fluxes to the coast; including: (1) the existence of multiple sources (fossil-fuel combustion from both mobile and stationary sources plus agricultural sources); (2) the difficulty in estimating dry deposition at the scale of whole watersheds; (3) the difficulty in measuring gaseous losses from ecosystems (including denitrification of nitrate to nitrogen as well as volatilization of ammonia and NOy compounds), and (4) the multiple pathways for nitrogen flows (surface waters, groundwaters, and atmosphere). (Generally, these problems are far less significant when dealing with phosphorus fluxes.) However, much insight into nitrogen fluxes to coastal waters has been gained recently by analyzing fluxes at relatively large spatial scales.

Over the past six years, the International SCOPE Nitrogen Project has been analyzing nitrogen fluxes at the scale of large regions, such as the combined watersheds of the North Sea, the combined watersheds of the northeastern United States from Maine through the Chesapeake Bay, and the Mississippi River basin. The International SCOPE Nitrogen Project was authorized by the International Council of Scientific Unions and has worked with other international efforts, including the International Geosphere-Biosphere Program, the United Nations’ Environmental Program, and the World Meteorological Organization. The motivation was to see what insights on nitrogen pollution could be gained by studying nitrogen biogeochemistry at a scale smaller than global but larger than small watersheds (Howarth 1996; Townsend 1999).

As one of its first activities, the International SCOPE Nitrogen Project evaluated nitrogen exports to the North Atlantic Ocean from the terrestrial landscape (both in Europe and in America) at the scale of large

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

regions (Howarth et al. 1996). At this scale, it is possible to evaluate the net influences of all the processes acting in the region at the scale of individual fields, feedlots, cities, forests, wetlands, and rivers. Some nitrogen inputs are actually easier to estimate at large spatial scales than at the scale of small watersheds and fields. For instance, dry deposition of nitrogen is difficult to estimate at the scale of individual watersheds and can be variable in space. However, dry deposition at the scale of large regions can be estimated with greater accuracy by using mass-balance constraints and knowledge of broad-scale atmospheric transport (Howarth et al. 1996; Prospero et al. 1996; Holland et al. 1999).

The International SCOPE Nitrogen Project showed large variations in the export of nitrogen to the North Atlantic from regions in the temperate zone, with fluxes per area of watershed varying from as low as 76 kg N km−2 yr−1 for the watersheds of northern Canada to 1,450 kg N km−2 yr1 for the watersheds of the North Sea (Figure 5-12). As stated above, the export of nitrogen from nonpoint sources dominates over wastewater and other point sources for all regions (Howarth et al. 1996; Howarth 1998). The flux of nitrogen from a region per area of watershed—both the total flux and the flux from nonpoint sources—is weakly correlated with population density (Figure 5-13).

The International SCOPE Nitrogen Project constructed mass balances for reactive nitrogen under human control at the scale of large regions (Howarth et al. 1996). For imports, the analysis carried out by the International SCOPE Nitrogen Project considered application of nitrogen fertilizer, nitrogen fixation by agricultural crops, deposition from the atmosphere of oxidized forms of nitrogen (which are presumed to come primarily from fossil-fuel combustion in the temperate zone; Holland et al. 1999), and the import or export of nitrogen in food and animal feedstocks. Sewage was not considered a net source, since it is a recycling of nitrogen that was brought into a region for agricultural purposes or directly as nitrogen in food. Similarly, deposition of ammonium and organic nitrogen from the atmosphere were not considered as inputs, since these are largely a recycling of nitrogen volatilized into the atmosphere from agricultural sources within the same region (Howarth et al. 1996; Howarth 1998).

Surprisingly, the International SCOPE Nitrogen Project found that the export of nitrogen from the landscape to the coast in the temperate zones of North America and Europe is a linear function of the import of reactive nitrogen forms into the region by human activity (Figure 5-14; Howarth et al. 1996; Howarth 1998). On average for the temperate regions of North America and Europe, 20 percent of the nitrogen inputs under human control flow out of regions to coastal waters. The majority of the human-controlled nitrogen inputs are either denitrified or stored in the ecosystems in regions. Unfortunately, the nature of these sinks—includ-

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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FIGURE 5-12 Regional export of total nitrogen to the North Atlantic coast per area of watershed (kg N km−2 yr−1). (A) Total nitrogen fluxes in rivers and in sewage treatment plants. (B) Fluxes in rivers that only originate from nonpoint sources of nitrogen in the landscape (modified from Howarth 1998).

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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FIGURE 5-13 The relationship between population density and the export of nitrogen in rivers to the coast for temperate regions surrounding the North Atlantic Ocean. Each point represents one region. (A) The total nitrogen export from the region in rivers. (B) The flux of nitrogen from nonpoint sources in the region, independent of upstream sources (modified from Smith et al. 1997).

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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FIGURE 5-14 A comparison of human-controlled inputs of nitrogen to a region and nitrogen export from the region to the coast in rivers, for temperate regions surrounding the North Atlantic Ocean. Note that the export of nitrogen from a region is linearly related to the inputs of nitrogen to the region. The dashed lines refer to the 95 percent confidence limits around the regression line (solid line; modified from Howarth et al. 1996).

ing whether or not they will change with time—is poorly known (Howarth et al. 1996).

The International SCOPE Nitrogen Project further used regressions (Figure 5-15A-D) to suggest that deposition from fossil-fuel sources (NOy deposition) per unit mass introduced into the landscape is a better predictor of nitrogen export to coastal waters (r2 = 0.81) than is fertilizer application (r2 = 0.28; Howarth 1998). Furthermore, a simple multiple regression that used both NOy deposition and agricultural inputs (i.e., the sum of fertilizer, nitrogen fixation in agriculture, and net movements of nitrogen in foodstocks) was constructed to predict nitrogen export to the coast. The best overall fit was obtained by a curve where the NOy deposition term was seven times greater than the agricultural input term (Howarth et al. 1996). This suggests that, per unit mass, nitrogen from fossil fuel sources may contribute more to the nitrogen flux in rivers to the coast than do agricultural sources. Of course, in many areas the total inputs of nitrogen as fertilizer are far greater than are the inputs from NOy deposition. For example, in the Mississippi River basin the total inputs of nitrogen as fertilizer far exceed those from NOy deposition; consequently, agri-

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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FIGURE 5-15 Analyses carried out during the International SCOPE Nitrogen Project suggest that nitrogen from deposition from fossil fuel sources (NOy deposition) per unit mass introduced into the landscape is a better predictor of nitrogen export to coastal waters (r2 = 0.81) than fertilizer application (r2 = 0.28; modified from Howarth 1998).

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

culture is the greatest contributor to the nitrogen export from that basin (Goolsby et al. 1999).

The best regression fit for the export of nitrogen from nonpoint sources for the temperate regions of the North Atlantic Ocean results from using the sum of NOy deposition and ammonium deposition to predict nitrogen export (Figure 5-15A-D; r2 = 0.92; Howarth 1998). The ammonium deposition is strongly tied to livestock densities (Bouwman et al. 1997), which suggests that livestock wastes contribute disproportionately to the nitrogen pollution of surface waters by agriculture and, together with the fossil-fuel source, are often major factors in nitrogen export to the coastal oceans at the scale of large regions (Howarth 1998).

INSIGHTS FROM THE SPARROW MODEL APPLIED TO THE NATIONAL SCALE

Another useful large-scale approach to assessing sources of nitrogen and phosphorus in surface waters was taken by Smith et al. (1997). By applying the Spatially Referenced Regressions on Watersheds (SPARROW) model (Appendix D) to a set of data from 414 stations in the National Stream Quality Accounting Network, Smith et al. (1997) concluded that just over half of the streams and rivers in the United States probably have total phosphorus concentrations in excess of 0.1 mg l−1 (Figure 5-16). Furthermore, they concluded that livestock waste production is the single largest source of phosphorus contamination leading to elevated phosphorus concentrations nationally (Smith et al. 1997). Mean values for “land-water delivery factors” (the percent of the original source of phosphorus that actually reaches surface waters) were estimated as approximately 0.07 and 0.11, respectively, for phosphorus from fertilizer application and phosphorus from livestock wastes. That is, the analysis by Smith et al. (1997) suggested that per mass of phosphorus, phosphorus from livestock wastes was 50 percent more likely to be exported to surface waters. Note that these delivery factors are estimated as part of the model in determining the best fit between nutrient sources and concentrations.

With respect to nitrogen, Smith et al. (1997) concluded that much of the United States probably exports less than 500 kg N km−2 yr−1, but that export is probably much higher in much of the Mississippi River basin and in the watersheds of the northeastern United States (Figure 5-17). For the areas of export over 1,000 kg N km−2 yr−1, Smith et al. (1997) concluded that fertilizer was the largest source of nitrogen overall (48 percent), followed by atmospheric deposition (18 percent) and livestock wastes (15 percent). To some degree, this result is driven by the large area of the Mississippi River basin—this basin represents 41 percent of the area of the lower 48 states, and is a region where fertilizer application greatly exceeds

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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FIGURE 5-17 Predicted local total nitrogen yield in hydrologic cataloging units of the conterminous United States. Local yield refers to transport per unit area at the outflow of the unit due to nitrogen sources in the unit, independent of upstream sources (Smith et al. 1997).

FIGURE 5-16 Classification of predicted total phosphorus concentrations in surface waters of the United States as estimated from the SPARROW model (Smith et al. 1997).

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

NOy deposition (Howarth et al. 1996). Conversely, in the northeastern United States, atmospheric deposition is the largest nonpoint source of nitrogen to surface waters (Howarth et al. 1996; Jaworski et al. 1997; Smith et al. 1997).

For the portions of the United States where total nitrogen export was over 1,000 kg N km−2 yr−1, the SPARROW model estimated land-water delivery factors of 0.24 for livestock wastes, 0.32 for fertilizer application, and 1.62 for atmospheric deposition (Smith et al. 1997). Note that for both livestock waste and fertilizer, the delivery factors are greater for nitrogen than for phosphorus (by two- to four-fold). This is consistent with the known greater mobility of nitrogen in dissolved forms in surface and groundwater and in volatile forms in the atmosphere.

There are some biases in the land delivery factors for nitrogen fertilizer and for atmospheric deposition, as Smith et al. (1997) did not include in their analysis the nitrogen fixation by agricultural crops or dry deposition from the atmosphere. Nitrogen fixation by agricultural crops tends to be correlated with nitrogen fertilizer application in the United States, and both are sources of nitrogen to downstream ecosystems (Howarth et al. 1996). Howarth et al. (1996) also demonstrated that, on average, for the portions of the United States that export over 1,000 kg N km−2 yr−1, nitrogen fertilizer makes up just over 60 percent of the sum of fertilizer application plus nitrogen fixation by agricultural crops. Adjusting the land-water delivery factor from Smith et al. (1997) to include nitrogen fixation as a source of nitrogen yields a new land-water delivery factor of 0.20 for the combined nitrogen from fertilizer and nitrogen fixation (a value comparable to that for nitrogen loss from livestock waste determined by Smith et al. 1997).

For atmospheric deposition, Smith et al. (1997) reported a land-water delivery coefficient of 1.62, suggesting that more nitrogen runs off the landscape from a depositional source than actually falls in deposition. This clearly cannot be so, and the most likely explanation for this high delivery factor is that the deposition estimates used for input were only for wet deposition of NOy, and did not include NOy dry deposition or wet or dry deposition of ammonium and organic nitrogen (personal communication, Smith 1999). On average for areas in the United States receiving fairly high levels of atmospheric deposition, wet NOy deposition is approximately 25 percent of total atmospheric deposition (wet and dry of both reduced and oxidized forms). (Although there is a great deal of uncertainty associated with this estimate [Johnston and Lindberg 1992; Lovett and Lindberg 1993; Whelpdale et al. 1997; Holland et al. 1999; Valigura et al. 2000]). Using this value as a correction factor for the land-water delivery factor for nitrogen deposition of Smith et al. (1997), leads

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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to a land delivery factor of approximately 0.40 in areas of high nitrogen export in the United States.

A comparison of these revised delivery factors of 0.40 for total atmospheric deposition of nitrogen (NOy) and 0.20 for nitrogen fertilizer application plus nitrogen fixation by agricultural crops leads to the conclusion that nitrogen from depositional sources is about two-fold more mobile in the landscape than is nitrogen running off agricultural fields. This conclusion is consistent with that from the regional analysis of the International SCOPE Nitrogen Project discussed earlier, which also demonstrated the greater mobility of nitrogen from NOy deposition (Howarth et al. 1996). Together, these results suggest that while the global mobilization of newly available nitrogen is greater through fertilizer production than through fossil-fuel combustion (Galloway et al. 1995; Vitousek et al. 1997), the nitrogen from fossil fuel sources may be disproportionately important to coastal eutrophication and other adverse impacts of nutrient over-enrichment.

Overall, the conclusions reached by Smith et al. (1997) from their SPARROW analysis agree remarkably well with the conclusions of the International SCOPE Nitrogen Project (Howarth et al. 1996; Howarth 1998), with one exception. Results from the Project show that livestock wastes are a more significant source of nitrogen to surface waters than predicted by Smith et al. (1997); the SPARROW analysis finds livestock wastes to be the major source of phosphorus, but a lesser source of nitrogen.

NUTRIENT BUDGETS FOR SPECIFIC ESTUARIES AND COASTAL WATERS

Knowledge of nutrient inputs to an estuary is essential for management of nutrient over-enrichment problems, and nutrient budgets have now been prepared for many estuaries. Several of these have recently been summarized by Valigura et al. (2000). Often, nutrient inputs are estimated as part of some larger scientific research project and are published in the peer-reviewed scientific literature. More frequently, the budgets are prepared as management tools and are either not published, or are published as government or consulting company reports (Valigura et al. 2000). Documentation of the data sources and approaches used is sometimes missing and is seldom fully adequate for independent review.

No standard methodologies exist for estimating nutrient inputs to estuaries, and many different approaches have been used. In some cases, nutrient budgets are based on export-coefficient models, where nutrient exports are estimated from literature values as a function of land-use types without independent verification of fluxes (Chapter 8). In other

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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cases, budgets are based on empirically derived loading coefficients for the actual watershed. These approaches work well for determining the importance of point-source inputs such as wastewater treatment plants. However, without proper calibration, estimates for nutrient inputs from non-point sources can be misleading. Estimating the importance of atmospheric deposition as a source is particularly problematic when using export-coefficient models.

For example, export-coefficient models simply take empirical data, and apply it through series of relatively straightforward calculations to obtain an estimate of the total load. In the simplest form (which is often the form used), the approach uses published coefficients for various land use types in the watershed (developing these coefficients is not straightforward, thus often the coefficients were derived for regions other than that within which the watershed resides). In a simple hypothetical watershed, published coefficients might suggest that farmland exports X g N m−2 yr−1, forests export Y g N m−2 yr−1, and urban lands export Z g N m2 year−1. These values are multiplied by the area of each land type in the watershed to get the export for the watershed as a whole. Atmospheric deposition of nitrogen (NOy) presents an immediate problem in that these models have historically not worried about whether the export coefficients used were derived for areas with high or low atmospheric deposition of nitrogen. Thus, atmospheric deposition has been ignored, and so the export from forests is generally treated as a background, natural flux. This erroneously implies that no amount of atmospheric deposition of nitrogen will increase the export of nitrogen from forests. Presumably, the approach could be improved so that forest export varied depending on deposition, but to date, no specific efforts to address this problem have been successfully completed.

Almost all nutrient budgets for estuaries rely on gauged stream discharge data where these are available. However, for many estuaries (including major ones such as Chesapeake Bay, Delaware Bay, and the Hudson River), significant portions of the watersheds are not gauged because of the difficulty in gauging tidal streams and rivers (Valigura et al. 2000). Where available, data on concentrations of total phosphorus and nitrogen are used in these budgets, but for many estuaries only inorganic dissolved nutrients are measured (Valigura et al. 2000). These problems add considerable error to the nutrient budgets.

Methodologies for determining the sources of nutrients and the magnitude of the load contributed by each are poorly developed at the scale of individual estuaries, and there is an urgent need for developing better approaches, particularly with regard to atmospheric deposition of nitrogen onto the landscape. The large-scale and regional analyses discussed above (the International SCOPE Nitrogen Project and the SPARROW

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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analysis) provide a potential framework based on quantifying inputs to the watershed, but these analyses are relatively recent and have not yet been applied to the management of most estuaries. In an effort to determine the validity of using SPARROW-derived estimates for a given estuary, Valigura et al. (2000) conducted a preliminary comparison of SPARROW-derived estimates with independently derived estimates of nitrogen loading to 27 estuaries on the Atlantic and Gulf of Mexico coasts of the United States. Based on that comparison, Valigura et al. (2000) concluded that while SPARROW accurately predicted the mean loading to the estuaries as a group, it did a poor job of predicting the load to any one particular estuary (i.e., a linear regression of the SPARROW estimates and the locally derived estimates had a slope of 1 and an R2 of 0.49). However, as with many such analyses involving locally derived information, the observed data from each estuary varies in quality and quantity and the methods used to calculate estimates varied as well. Thus, the locally derived estimates were not obtained from directly comparable data sets and most were not verified. Thus the poor match between SPARROW predictions and local estimates may lie with the quality of the individual estimates for the 27 estuaries. (Chapters 7 and 8 expand on the limitations imposed on understanding individual estuarine behavior by inconsistent observations.)

Perhaps the greatest uncertainty with estuary nitrogen budgets concerns the contribution of atmospheric deposition. In most classical estuarine studies, nitrogen inputs from the atmosphere were completely ignored. This has changed since Fisher and Oppenheimer (1991) pointed out the potential importance of atmospheric deposition as a source of nitrogen to Chesapeake Bay, and since Paerl (1985) showed the importance of atmospheric deposition as a nitrogen source to the coastal waters of North Carolina. However, even many nutrient budgets constructed during the last decade have no estimate for the input of nitrogen from atmospheric deposition. In many other estuaries, budgets estimate the importance only of direct deposition onto the surface waters of the estuary itself (and generally only wet deposition, not dry deposition), and do not estimate deposition onto the landscape with subsequent export to the estuary.

Available evidence (although constrained by limited monitoring) indicates that direct deposition onto the water surface alone (not including the contribution of nitrogen which falls on the landscape and is then exported to estuaries) contributes between 1 percent and 40 percent of the total nitrogen input to an estuary—depending in large part on the relative area of the estuary and its watershed (Nixon et al. 1996; Valigura et al. 2000). In estuaries where the ratio of the area of the estuary to the area of its watershed is greater than 0.2, direct atmospheric depositions usually

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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make up 20 percent or more of the total nitrogen loading (Valigura et al. 2000). Where the ratio of the estuarine area to the area of its watershed is less than 0.1, atmospheric deposition directly onto the water surface generally makes up less than 10 percent of the total nitrogen input (Valigura et al. 2000).

For estuaries that have relatively large watersheds, the deposition of nitrogen from the atmosphere onto the landscape with subsequent runoff into the estuary is probably greater than the deposition of nitrogen directly onto the water surface. Unfortunately, the magnitude of this flux is poorly characterized for most estuaries. The deposition onto the landscape can be estimated for most watersheds, although the error associated with these estimates can be considerable due to inadequate monitoring and the difficulty with measuring dry deposition. The larger problem, however, is with determining what portion of the nitrogen deposition is retained in the landscape and what portion is exported to rivers and the coast. The two major approaches for making this determination are to use statistical models or to use process-based models on nitrogen retention in the watershed. In their application to estuaries, both approaches are quite recent and are relatively untested. There is an urgent need for further development and evaluation of these techniques; however, it appears that the statistical approaches have led to more reliable estimates, for reasons discussed below.

Both the SPARROW model and regressions comparing nitrogen flux in rivers to sources of nitrogen across landscapes (used by the International SCOPE Nitrogen Project) represent examples of statistical approaches that appear to provide reliable estimates of the portion of the nitrogen deposition retained in the landscape versus what is exported to rivers and coastal areas. Jaworski et al. (1997) used a similar approach in the northeastern United States, comparing atmospheric deposition and riverine flux for 17 watersheds with relatively little agricultural activity or sewage inputs. This led to the conclusion that approximately 40 percent of the nitrogen deposition is exported from the landscape (correcting their analysis by assuming that dry deposition is equal to wet deposition), a value remarkably similar to the results from applying the SPARROW model at the national scale. By applying this result to other watersheds in the northeast, including those with agricultural activity, Jaworski et al. (1997) estimated that between 36 percent and 80 percent of the total nitrogen flux in rivers was originally derived from atmospheric deposition onto the landscape. Note that the riverine nitrogen fluxes were estimated at U.S. Geological Survey (USGS) gauging stations above the tidal portions of these rivers, and generally excluded the large urban influences at the river mouths.

In another recent effort, a National Oceanic and Atmospheric Admin-

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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istration (NOAA)-sponsored project brought together researchers from around the United States to examine atmospheric deposition to coastal waters (Valigura et al. 2000). Valigura et al. (2000) summarized and compared the four different approaches included in the NOAA project, including a process-based model and an application of the statistical approach used by SPARROW. They report that, for 42 estuaries in the United States, atmospheric deposition onto the landscape contributed between 6 percent and 50 percent of the total nitrogen load to the receiving body. Jaworski et al (1997) and Valigura et al. (2000) give estimates in common for only one river/estuary—the Hudson-Raritan—and for this system, their estimates are similar to the statistical model results, but quite different from the process-based model estimates. Jaworski et al. (1997) estimate that 34 percent of the nitrogen flux in the Hudson comes from atmospheric deposition onto the landscape, after correction for the point source inputs from New York City (Hetling et al. 1996). In contrast, estimates from the process-based model indicated 9 percent of the nitrogen flux of the Hudson-Raritan total nitrogen load comes from nitrogen deposition onto the landscape. The statistical SPARROW model approach estimated the flux to the estuary from atmospheric deposition onto the watershed as 26 percent for this system.

Great uncertainty about the importance of atmospheric deposition as a nitrogen source to specific estuaries may exist. However, there is little doubt that the relative importance of fossil-fuel combustion versus agricultural activity in controlling atmospheric deposition of nitrogen to estuaries depends both on the nature and extent of farming activities in the watershed and on the nature and extent of fossil-fuel combustion in the airsheds upwind of the watershed. In estuaries fed by watersheds with little agricultural activity but significant loads of atmospheric pollution (such as the Connecticut and Merrimack rivers and most of the northeastern United States), atmospheric deposition of nitrogen from fossil-fuel combustion can account for up to 90 percent or more of the nitrogen contributed by nonpoint sources. On the other hand, for watersheds such as the Mississippi Basin where agricultural activity is high and atmospheric pollution from fossil-fuel combustion is relatively low (Figure 5-18), agricultural sources dominate the fluxes of nitrogen. Interestingly, the major hot-spots of agricultural activity that dominate the nitrogen fluxes for the Mississippi and Gulf of Mexico appear to be far from the Gulf in Iowa, Illinois, Indiana, Minnesota, and Ohio (Goolsby et al. 1999).

For many estuaries, both atmospheric deposition of nitrogen derived from fossil-fuel combustion (NOy) and nitrogen from agricultural sources are likely to be major contributors. For example, the model used by managers to estimate nitrogen inputs to Chesapeake Bay predicts that

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

FIGURE 5-18 A comparison of human-controlled inputs of nitrogen and nitrogen losses (kg N km−2 yr−1) as food exports and in riverine exports between the northeastern United States and the Mississippi River basin. Note that, on average, nitrogen is exported in foods and feedstocks from the Mississippi basin and imported to the northeastern United States.

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

agriculture contributes 59 percent of the nonpoint source inputs, NOy deposition onto the landscape is slightly less important (Magnien et al. 1995). The comparative analysis of Jaworski et al. (1997), on the other hand, suggests that atmospheric deposition is the dominant source of nitrogen from nonpoint sources in the major tributaries of Chesapeake Bay. Further study and analysis is necessary to determine whether Jaworski et al. (1997) have overestimated the importance of atmospheric deposition or whether Magnien et al. (1995) have underestimated it.

However, the process-based model of nitrogen retention used by Magnien et al. (1995) has not been independently verified and is subject to large uncertainties (Boesch et al. 2000). Small changes in the assumed ability of forests to retain or export nitrogen from atmospheric deposition can lead to large changes in the relative importance of NOy deposition to the bay. As discussed above, there is great variation among forests in their ability to retain nitrogen from atmospheric deposition, and regional and large-scale analysis of nitrogen fluxes for the United States indicate a greater mobility of nitrogen from deposition (less retention) than is often found in small-scale watershed studies. Further, the model of Magnien et al. (1995) does not include some of the latest findings on nitrogen export from land, such as the large export of nitrogen in dissolved organic forms that was noted above (Campbell et al. 2000).

A recent report from the Environmental Protection Agency (EPA) estimates that between 10 percent and 40 percent of the total nitrogen input to estuaries comes from atmospheric deposition, including deposition directly onto the water surface and onto the watershed (EPA 1999c). However, it must be stressed that very few of the individual studies upon which this conclusion is based had adequate methodologies for determining the input of nitrogen from atmospheric deposition, particularly the indirect input through atmospheric deposition onto the landscape with subsequent runoff into the estuary. Many of these studies have probably underestimated the importance of this pathway, and it seems likely that atmospheric deposition is a greater input to estuaries than suggested by the 1999 EPA report.

OCEANIC WATERS AS A NUTRIENT SOURCE TO ESTUARIES AND COASTAL WATERS

In addition to receiving nutrient inputs from land and from atmospheric deposition, estuaries can receive nutrients across their boundary with the ocean. This term is often ignored, but can be substantial. For example, Nixon et al. (1995) estimate that for total nutrient inputs to Narragansett Bay, 15 percent of the nitrogen and 40 percent of the phosphorus inputs come from offshore, oceanic sources; despite this, the net

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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flux of both nitrogen and phosphorus for Narragansett Bay is an export of these nutrients from the estuary to offshore waters (Nixon et al. 1996). On the other hand, Chesapeake Bay is a net importer of phosphorus from offshore ocean waters, although it too is a net exporter of nitrogen (Boynton et al. 1995; Nixon et al. 1996). The physical circulation pattern of an estuary is a major determinant in the importance of nutrient import to the estuary from offshore sources. Partially mixed estuaries (such as Chesapeake Bay) and fully mixed estuaries (such as Narragansett Bay) often import nutrients from offshore, whereas salt-wedge estuaries (such as the southwest pass of the Mississippi River and Oslo Fjiord) and hypersaline estuaries (such as portions of Shark Bay, Australia) do not (Howarth et al. 1995).

Offshore waters on the continental shelf can themselves receive nutrients from several sources, including deep ocean water, river and sewage inputs from land, and direct deposition from the atmosphere (Nixon et al. 1996; Prospero et al. 1996; Howarth 1998). The relative importance of these sources varies among the coastal waters of the United States, in part because of differences in ocean circulation patterns (particularly advection of water from the deep ocean—water that is extremely high in nutrients—onto the continental shelf). For most of the continental shelf area of the United States, this advection of water is the dominant nutrient input. However, input from the Mississippi River is the dominant source for the Gulf of Mexico. Human activity has tended to greatly increase inputs of nitrogen from rivers and atmospheric deposition, but has had no impact on the advection of water from the deep ocean onto the continental shelf. Consequently, human activity has almost tripled nitrogen input to the Gulf of Mexico, but has increased nitrogen inputs to the waters on the continental shelf of the northeastern United States by only 28 percent (Table 5-1). Of course, much of this input in the northeastern United States is concentrated in the plumes of a few rivers, such as that of the Hudson River, and these waters may therefore be experiencing eutrophication (Howarth 1998).

Rate of Change of Nutrient Inputs to the Coast

Historical data on fluxes of total nitrogen in rivers are rare, but data for trends in nitrate concentrations are available for many rivers going back to the early 1900s. Since human activity preferentially mobilizes nitrate over other forms of nitrogen in rivers (Howarth et al. 1996), these historical nitrate data are valuable in tracking the effects of humans on nitrogen fluxes to the coast. For the Mississippi River, the nitrate flux to the Gulf of Mexico is now some three-fold larger than 30 years ago, and most of this increase occurred between 1970 and 1983 (Figure 5-19;

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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TABLE 5-1

 

Nitrogen Sources (Tg yr-1)

North Atlantic Continental Shelves

Rivers and Estuaries

Direct Atmospheric Deposition

Deep Ocean

Increase Due to Humans (%)

North Canada rivers

0.16 (0.16)

0.10 (0.03)

0.77

7

St. Lawrence basin

0.34 (0.11)

0.13 (0.01)

1.26

25

Northeast coast of the United States

0.27 (0.03)

0.21 (0.01)

1.54

28

Southeast coast of the United States

0.13 (0.03)

0.06 (0.01)

1.36

11

Gulf of Mexico

2.10 (0.50)

0.28 (0.03)

0.14

275

North Sea and Northwest Europe

0.97 (0.14)

0.64 (0.02)

1.32

98

Southwest European coast

0.11 (0.04)

0.03 (0.001)

0.20

40

TABLE 5-1 Sources of nitrogen to the continental shelves of the temperate zone portions of the North Atlantic Ocean. Flux from rivers and estuaries is the direct input of rivers that discharge onto the continental shelf, minus nitrogen consumed in estuaries. Atmospheric deposition estimates are those directly onto the waters of the continental shelf and do not include deposition onto the landscape (which is part of the flux from rivers and estuaries). The flux from the deep ocean represents the advection of nitrate-rich deep Atlantic water onto the continental shelf. Data for modern values are means reported by Nixon et al. (1996). Pristine values as outlined by Nixon et al. (1996) for their treatment of modern estimates, but with data for pristine river fluxes from Howarth et al. (1996) and for pristine values of deposition from Prospero et al. (1996). “Increase due to humans” is the percentage comparison of total modern inputs compared to pristine inputs. Fluxes from the deep ocean are assumed not to have been affected by human activities (modified from Howarth 1998).

Goolsby et al. 1999). Similarly, nitrate fluxes in many rivers in the northeastern United States have increased two- to three-fold or more since 1960, with much of this increase occurring between 1965 and 1980 (Figure 5-20; Jaworski et al. 1997). Interestingly, most of the increase in nitrate in the Mississippi River was due to increased use of nitrogen fertilizer (Goolsby et al. 1999), whereas most of the increase in nitrate in the northeastern rivers was due to increased nitrogen deposition from the atmosphere onto the landscape, with the nitrogen originating from fossil-fuel combustion (Jaworski et al. 1997). The increase in nitrate flux in the

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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FIGURE 5-19 Bar chart showing the annual flux of nitrogen as nitrate (NO3) from the Mississippi River basin to the Gulf of Mexico, indicating significant increases beginning in the late 1970s (modified from Goolsby et al. 1999).

FIGURE 5-20 Flux of nitrate nitrogen from five major rivers in the northeastern United States from the early 1900s to 1994 (modified from Jaworski et al. 1997).

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

northeastern rivers during the 1960s and 1970s, and its stabilization since then, closely parallels the trend in human inputs of nitrogen to the landscape during that time (Jaworski et al. 1997).

In contrast to nitrogen, phosphorus fluxes to estuaries have often changed little over the past several decades. For the Mississippi River, data on total phosphorus flux are only available since the early 1970s, but there has been no statistically significant change since then (Goolsby et al. 1999). Smith et al. (1987) used data from 300 river locations throughout the United States to compare water quality trends from 1974 to 1981. Many rivers showed no trend during that time; rivers that had a trend in total phosphorus flux were equally divided between those that showed an increase and those that showed a decrease. Where total phosphorus fluxes increased, it was generally attributable to increased use of phosphorus fertilizer in the watershed. Decreases in total phosphorus fluxes were generally a result of point source reductions (Smith et al. 1987). Smith et al. (1987) also analyzed the national river data for trends in nitrate flux from 1974 to 1981. For nitrate, most rivers showed a marked increase in flux during that time, particularly for rivers in the eastern United States. This increased nitrate flux was attributed both to agricultural activity and to nitrogen deposition (Smith et al. 1987).

IMPLICATIONS FOR ACHIEVING SOURCE REDUCTIONS

Human activity has an enormous impact on the cycling of nutrients and especially on the movement of such nutrients as nitrogen and phosphorus into estuaries and other coastal waters. Although much effort has been made in the United States to improve control of point sources of pollution, nonpoint sources as urban runoff, agricultural runoff (particularly from animal feeding operations), and atmospheric deposition are generally of greater concern in terms of impact on nutrient enrichment and eutrophication of coastal waters. While sewage inputs dominate in some estuaries, nonpoint sources dominate nationally. Insufficient effort has been expended on controlling nonpoint sources of nitrogen and phosphorus, and there are few comprehensive plans for managing nutrient enrichment of the nation’s coastal waters, particularly from nonpoint sources. Efforts to manage nonpoint and point sources of nitrogen and phosphorus are needed to reduce adverse impacts of nutrient over-enrichment in the nation’s rivers, lakes, and coastal waters.

There is evidence that both atmospheric deposition of nitrogen from fossil-fuel combustion and agricultural sources of nitrogen contribute nitrogen to coastal waters. The relative importance of these varies among estuaries, but recent evidence indicates that the amount of nitrogen from

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

deposition has been historically underestimated as an input to many estuaries, particularly by the indirect pathway of nitrogen deposited onto the landscape and then exported to the estuary. Recent evidence also indicates that per unit input to the landscape, nitrogen from fossil-fuel combustion is more important than nitrogen from fertilizer and, in turn, contributes disproportionately in the input of nitrogen to coastal waters.

Much uncertainly remains regarding the fluxes of nitrogen from the atmosphere to the landscape and to estuaries, and this is a critically important research priority. Although understanding some details regarding the atmospheric transport and fate of biologically available nitrogen will require additional research, the significant role atmospheric deposition of nitrogen plays in nutrient over-enrichment in some regions is clear. Addressing this component of the problem will require coordinated efforts over many states, clearly dictating a federal role in the effort. The regional nature of the atmospheric component of nitrogen loading argues that nutrient management should be a significant component of efforts to reduce air pollution and should be a key consideration during re-authorization of the Clean Air Act.

In general, sources of nutrients to estuaries have been poorly characterized, and in some cases sources have been mistakenly characterized because some land-use export-coefficient models used for characterization are inadequately verified. There are currently no easy-to-use and reliable methods for the manager of an estuary to determine the sources of nutrients flowing into that estuary. As will be discussed in Chapter 8, enhanced and coordinated monitoring efforts will be a key component of any local, regional, or national effort to reduce the impacts of nutrient over-enrichment.

Some critical questions related to understanding the sources of nutrients most affecting eutrophication and other impacts of nutrient over-enrichment remain unanswered. For instance, nitrogen deposition and fate in urban and suburban areas is poorly known, and wet nitrogen deposition in coastal areas is poorly understood. There is only a limited understanding of dry deposition in any environment, and understanding this in coastal areas and over water is challenging. Research efforts to expand understanding of atmospheric deposition of nitrogen should be expanded.

Changes in agricultural production systems are concentrating large amounts of nutrients in localized areas, thereby increasing the risk of nutrient leakage to the environment. Most of this concentration is associated with animal feedlots and with the long-distance transport of feedstocks. Changes in farm practices are driven by economics, and this concentration and long-range transport provide economic advantages to

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
×

the producers; the larger costs, such as the external cost of nutrient exports to estuaries, remain unaddressed. As is discussed further in Chapter 9, a balanced and cost-effective nutrient management strategy will require an understanding of both the relative importance of various sources of nutrients, and the economic costs associated with reducing the loads attributable to each.

Suggested Citation:"5 Sources of Nutrient Inputs to Estuaries and Coastal Waters." National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. Washington, DC: The National Academies Press. doi: 10.17226/9812.
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Environmental problems in coastal ecosystems can sometimes be attributed to excess nutrients flowing from upstream watersheds into estuarine settings. This nutrient over-enrichment can result in toxic algal blooms, shellfish poisoning, coral reef destruction, and other harmful outcomes. All U.S. coasts show signs of nutrient over-enrichment, and scientists predict worsening problems in the years ahead.

Clean Coastal Waters explains technical aspects of nutrient over-enrichment and proposes both immediate local action by coastal managers and a longer-term national strategy incorporating policy design, classification of affected sites, law and regulation, coordination, and communication.

Highlighting the Gulf of Mexico's "Dead Zone," the Pfiesteria outbreak in a tributary of Chesapeake Bay, and other cases, the book explains how nutrients work in the environment, why nitrogen is important, how enrichment turns into over-enrichment, and why some environments are especially susceptible. Economic as well as ecological impacts are examined.

In addressing abatement strategies, the committee discusses the importance of monitoring sites, developing useful models of over-enrichment, and setting water quality goals. The book also reviews voluntary programs, mandatory controls, tax incentives, and other policy options for reducing the flow of nutrients from agricultural operations and other sources.

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