Natural Resources Conservation Service

Fate and Transport of Nutrients: Nitrogen

Working Paper No. 7

Ronald F. Follett
USDA, Agricultural Research Service
Soil-Plant-Nutrient Research Unit
Fort Collins, Colorado

September 1995

Contents

Abstract
The Nitrogen Cycle
Fate and Transport Processes
Effect of Nitrogen Source on Fate and Transport
Primary and Secondary Flows of Nitrogen
Concerns about Nitrogen's Impacts on the Environment
Nitrogen Use Efficiency
Literature Cited

Abstract

Nitrogen (N) is ubiquitous in the environment. It is also one of the most important nutrients and is central to the production of all crop plants. However, N also forms some of the most mobile compounds in the soil-plant-atmosphere system. There is mounting concern about agriculture's role in N delivery into the environment, as nitrogen represents the mineral fertilizer most applied to agricultural land. This is because available soil-N supplies are generally inadequate for optimum crop production. This manuscript reviews the fate and transport of N from the various sources used to supply the N-requirements of crops in the context of the N cycle. Use of N budgets, or a mass balance approach, is needed to understand the options for improving management of N in farming and livestock systems and for mitigating the environmental impacts of N. Fertilizing crops for crop-N uptake that will be near the point of maximum yield is in general an economically and environmentally acceptable practice. The management objective is to lower the rate and duration of the loss processes themselves. Practices and concepts are considered that lessen the opportunity for loss processes to occur and that help to decrease the amount of N that may be lost to the environment. In some cases improved efficiency is achieved by using less nutrients; in other cases it can be achieved by increasing the yield while using the same amount of N-input. In either case, the goal is to decrease the total residual mass of N in the soil. Another approach is to keep the residual N within the soil-crop system by curtailing the transport processes (leaching, runoff, erosion, and gaseous losses) that carry pollutants out of the system.

Nitrogen

Nitrogen (N) is ubiquitous in the environment. It is one of the most important nutrients and is required for the survival of all living things. It is also central to the production of all crop plants. Nitrogen accounts for 78 percent of the atmosphere as elemental dinitrogen (N₂) gas. In the form of dinitrogen (N₂) gas, N is inert, does not impact environmental quality, and is not directly available for plant uptake and metabolism. However, N also forms some of the most mobile compounds in the soil-plant-atmosphere system. The mounting concerns related to agriculture's role in N delivery into the environment are reflected in several detailed reviews (Follett 1989; Follett et al. 1991; Hallberg 1987, 1989; Keeney 1986; Power and Schepers 1989). Nitrogen represents the mineral fertilizer most applied to agricultural land, because available soil-N supplies are generally inadequate for optimum crop production. The fate and transport of N from the various sources used to supply the N-requirements of crops must be considered in the context of the N cycle. Use of N budgets, or a mass balance approach, is needed to understand the options for improving management of N in farming and livestock systems and for mitigating the environmental impacts of N.

The Nitrogen Cycle

Biological Nitrogen Fixation

Through the process of biological N-fixation (BNF), symbiotic (mutually beneficial) and nonsymbiotic organisms can fix atmospheric N₂ gas into organic N forms (Figure 1). A few living organisms are able to utilize molecular N₂ gas from the atmosphere. The best known of these are the symbiotic Rhizobia ("legume bacteria"), nonsymbiotic free-living bacteria such as Azotobacter and Clostridium, and the cyanobacteria. Generally, in a symbiotic relationship, one organism contains chlorophyll and uses light energy to produce carbohydrates. The other organism receives some of the carbohydrates and uses them as an energy source to enzymatically fix atmospheric N₂ into the ammonia (NH₃) form of N and thence into amino acids and other nitrogenous compounds that are nutritionally useful to the chlorophyll-containing organism. To agriculture, the most important type of BNF is symbiotic fixation by legumes (i.e. alfalfa, clovers, peas, beans, etc.). Follett et al. (1987) have estimated that in the United States about 1.5 million pounds of symbiotically fixed-N are returned annually to cropland soils from leguminous crops.

Mobilization and Immobilization of Soil Nitrogen

Nitrogen taken up by plants from the soil originates from indigenous organic or inorganic forms. Organic N occurs naturally as part of the soil's organic matter fraction; it can also be added to the soil from manure, symbiotic and nonsymbiotic biological N-fixation, plant residues, and from other sources. Soil microorganisms and their activities are an integral part of mineralization and immobilization processes in soil (Figure 1); soil-organic N can be transformed to ammonium (NH₄⁺) by the process of ammonification. Inorganic (mineral) forms of N include NH₄⁺ and nitrate (NO₃^-), both readily taken up by crops, and nitrite (NO₂^-) that occurs as an intermediate form during mineralization of NH₄⁺ to NO₃^-. Usually NO₂^- does not accumulate in soils because it is rapidly transformed to NO₃^- or denitrified.

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Figure 1. The nitrogen cycle
Immobilization of NO₂^- and NO₃^- back to organic forms of N can also occur through enzymatic activities associated with plant or microbial N-uptake and N-utilization processes.

Nitrate, a water-soluble anion, is not sorbed to the negatively charged sites on soil colloids (the cation exchange capacity or CEC of the soil), is very mobile, and moves readily with percolating water (leaching). Ammonium, a cation, can be sorbed to the CEC, incorporated (fixed) into clay and other complexes within the soil, released by weathering back into the available mineral pool, or immobilized into organic form by soil microbial processes. Ammonium that is associated with soil colloids can be transported into surface water during water or wind erosion of soil or, under certain conditions, can volatilize into the atmosphere as NH₃ gas and be aerially transported across the landscape including into surface water. Gaseous NH₃ often is returned to the soil-plant system by direct uptake into plant leaves or dissolved in precipitation.

Gaseous Losses of Nitrogen to the Atmosphere

Ammonia volatilization--Ammonium ions in the soil solution enter into an equilibrium reaction with NH₃ in the soil solution. The soil solution NH₃ is, in turn, subject to gaseous loss to the atmosphere. Soil pH and concentration of NH₄⁺ in the soil solution are important factors affecting amount of NH₃ loss to the atmosphere. As soil pH increases, the fraction of soil-solution NH₄⁺ plus soil-solution NH₃ in the NH₃ form also increases by an order of magnitude for every unit of pH above 6.0, thus increasing losses of soil-solution NH₃ to the atmosphere. As summarized by Stevenson (1986), NH₃ volatilization:

• is of most importance on calcareous soils, especially as soil pH exceeds 7;
• losses increase with temperature and can be appreciable for neutral or alkaline soils as they dry out;
• is greater in soils of low CEC, such as sands;
• losses can be high when high-N organic wastes, such as manure, are permitted to decompose on the soil surface;
• losses are high from urea applied to grass or pasture, as a result of hydrolysis of the urea to NH₃ by indigenous urease enzyme;
• losses from soil and fertilizer N are decreased by growing plants.

Denitrification--As organic matter in soil decomposes, first NH₄⁺, then NO₂^- and finally NO₃^- ions are formed by the process of nitrification (Figure 1). Nitrite usually does not accumulate in soils because it is rapidly transformed to NO₃^- or is denitrified to N₂ gas, nitrous oxide (N₂O), nitric oxide (NO), or one of the other gaseous nitrogen oxide (NO_x) compounds. Nitrate can also be lost to the atmosphere through the denitrification processes. Nitrous oxide, a product of incomplete denitrification, is a greenhouse gas and may contribute to global climate change and to thinning of the ozone layer.

 

Not only denitrification (a reductive process), but also the oxidative process of nitrification causes emission of a small amount of NO₂ (Tortoso and Hutchinson 1990). However, denitrification is the route for most losses of gaseous N compounds to the atmosphere. The potential for denitrification is increased as oxygen levels in the soil decrease. Nitrate, a desirable nutrient for plants, is reduced first to NO₂^-, then to NO, next to N₂O, and finally to N₂. In addition to limited oxygen, denitrification also generally needs effective microbes, suitable reducing agents like organic C, and oxides of N (Firestone and Davidson 1989, Klemedtsson et al. 1988).

Animal/Plant Nitrogen Uptake and Cycling

Nitrogen cycling for pasture systems is also shown in Figure 1. Inputs are from fertilizers and manures, biological N-fixation, wet and dry deposition from the atmosphere, supplemental feed to livestock, and mineralization of soil organic matter. Losses may occur through harvest of animal or plant products, transfer of N within the pasture with animal excreta, fixation of N in the soil, volatilization, leaching, soil erosion, and surface runoff. The soil compartment includes a pool of available N (NO₃^- and NH₄⁺) for plant uptake that is in equilibrium with N in residues (organic-N) and, especially for some soils, with fixed NH₄⁺ which is held between mineral layers of the clay. Plant N-uptake is from the available soil pool. The N in the herbage is eaten by grazing animals and either utilized by the animal or excreted as feces or urine and returned to the soil. Nitrogen is released from excreta and plant residues to the available nutrient pool in the soil.

Role of soil organisms--Soil microfauna and microflora have a major role in N cycling. Release of N from plant and animal residue depends on microbial activity. Soil bacteria utilize the more readily available soluble or degradable organic fractions. Fungi and actinomycetes decompose the resistant cellulose, hemicellulose, and lignin. Dung beetles, earthworms, and other soil fauna increase the decomposition rates of feces and plant litter by mixing them with soil. Rhizobia and vesicular arbuscular mycorrhizae (VAM) associate with plant roots to fix N and increase nutrient- and water-scavenging ability, respectively. VAM infection of roots is considered more helpful for tap-rooted pasture legume species than for fibrous-rooted grasses. At any time, soil-microbial biomass contains much of the actively cycling N of the soil and represents a relatively available N-pool, capable of rapid turnover (Bristow and Jarvis 1991).

Role of the grazing animal--Grazing animals affect plant growth by defoliation, traffic patterns, herbage fouling, partitioning of ingested N to body weight, feces, and urine, redistribution of herbage N in excreta, and N-turnover rate. Defoliation by grazing animals prevents senescence of plant tissue, removes N in animal products, changes the N pathway from internal plant recycling or leaf fall to return as feces and urine, increases light penetration into the canopy and, through selective grazing, may alter botanical composition by promoting one species over another. Animal traffic compacts soil, sometimes making soil characteristics for plant growth less desirable. Herbage fouling by feces reduces its acceptability for grazing, thereby increasing maturity and reducing forage quality and/or consumption by grazers. Urine does not cause herbage to be unacceptable for grazing.

Animals on range may utilize more of the forage near watering points. Greater density of dung and increased levels of soil-profile NO₃^- are frequently observed in areas near watering and shade points (Haynes and Williams 1993, Wilkinson et al. 1989). Even without transfer of N to unproductive areas such as woods, shade, watering points, fence lines, and cow paths, the consumption and excretion habits of ruminants have the result that N is gathered from large areas of the pasture and returned to smaller areas. This concentrating effect frequently means that N cycled through livestock cannot be used efficiently by forage plants. On an annual basis, less than 35 percent of pasture areas receive excretal N; some areas receive one or more application (overlapping of excreta). This uneven distribution means some of the pasture will be underfertilized, and some overfertilized.

Fate and Transport Processes

Leaching

Nitrate (NO₃^-) is a negatively charged ion that is repelled by (rather than attracted to) negatively charged clay mineral surfaces in soil (the CEC). It is the primary form of N leached into groundwater, is totally soluble at the concentrations found in soil, and moves freely through most soils. As described by Jury and Nielson (1989), movement of the NO₃^- ion through soil is governed by convection, or mass-flow, with the moving soil solution and by diffusion within the soil solution. The widespread appearance of NO₃^- in ground water is a consequence of its high solubility, mobility, and easy displacement by water.

Most soil and environmental parameters which influence the transport of dissolved NO₃^- through natural field soils vary substantially at different locations, even at short separation distances. For example, a single wormhole or decayed root channel can raise the infiltration rate of a surface soil by orders of magnitude when water is ponded over it. Therefore, even when water is applied relatively uniformly over an entire soil surface, such as with rainfall or sprinkler irrigation; resulting vertical water velocities within the soil are not spatially uniform, even when plant roots are absent. Under ponded conditions, solute velocity variations are usually much greater than for unsaturated conditions.

Erosion

Movement of N into surface water primarily occurs with soil erosion by water, rather than by wind. Briefly, soil erosion by water includes the processes of detachment, transport, and deposition of the soil particles by raindrops or surface flow (Foster et al. 1985). Some sediments may travel only a few millimeters while others may be transported long distances before they are either deposited or reach a surface water resource (i.e., a lake, reservoir, or stream). Soil erosion is important to the movement of N into surface water. Nitrogen as ammonium (NH₄⁺) is sorbed to the surfaces of clays and finer sediments or to the soil organic matter and is in organic-N forms in the soil organic matter. Nitrogen that degrades surface water is primarily transported in soil organic matter or as NO₃^-, a form that is completely water soluble.

Detachment of sediments and nutrients from the parent soil is selective for soluble nutrients (such as NO₃^-) and for the fine soil fractions that the nutrients (such as NH₄⁺ and the soil organic matter-N) are associated with. Therefore, N contained in runoff and/or associated with sediments is present in higher concentrations than in the parent soil. This difference is termed the enrichment ratio (ER). Enrichment of sediment loads is a two-step process: enrichment during particle suspension and enrichment due to redeposition of coarser particles during overland and channel flows. In order for management practices to decrease the effect of water erosion processes on the production and transport of sediment-associated N, they must directly influence the processes involved. Practices that leave crop residues on the soil surface are increasingly being accepted as a means to protect against soil particle detachment, slow sediment transport, and enhance deposition.

However, much still needs to be learned about effects of residue management practices on nutrient transport from agricultural fields. Follett, et al. (1987) assessed effects of tillage practices and slope on amount of organic-N in eroded sediments from cultivated land surfaces in Minnesota for major land resource areas (MLRA's) 102, 103, 104, and 105. Their estimates, using the Universal Soil Loss Equation (USLE), average organic matter in topsoil by slope category, and dominating slope gradient and soil series, indicate that conservation tillage compared to conventional tillage decreases the amount of organic-N associated with eroded sediments by about half; some additional decrease is obtained by using no-tillage. One can assume that added fertilizer N, which is sorbed to clay surfaces and to finer sediments or to soil organic matter, responds the same.

Runoff

As in the case of N-leaching, amount and timing of rainfall and soil properties are key factors that influence loss of dissolved N in runoff. Soils with low runoff potential usually have high infiltration rates, even when wet. They often consist of deep, well drained to excessively drained sands or gravels. In contrast, soils with high runoff potential have one or more of the following characteristics: very slow infiltration rates when thoroughly wetted, high clay content possibly of high swelling potential, high water tables, a claypan or clay layer at or near the surface, or shallowness over nearly impervious material. A combination of soil conditions of high runoff potential with high amounts of precipitation is especially conducive to surface runoff losses. Steeper slope gradients increase amount and velocity of runoff, while depressions, soil roughness, and the presence of vegetative cover or crop residues decrease runoff by improving water infiltration.

The dissolved concentration of N in surface runoff from soils under conservation- or no-tillage often is higher than from soils under conventional tillage (McDowell and McGregor 1984; Romkens 1973). Reasons may include incomplete incorporation of surface-applied fertilizer, dissolved nutrient contributions from decaying crop residues, and higher dissolved N concentration in the surface soil because of residue accumulation and decomposition. In addition, high concentrations of soluble N can occur when the presence of a soil horizon barrier (e.g., a fragipan) in the soil profile results in return flow of leached N back to the soil surface (Lehman and Ahuja 1985, Smith et al. 1988).

Some of the interpretations associated with agricultural nutrient management for crop production and the use of conservation tillage for erosion control, as they affect dissolved nutrients in surface- and subsurface-water discharges, are illustrated by a 10-year study reported by Alberts and Spomer (1985). Their study site was in the deep loess hills in western Iowa. The loess is underlain by nearly impervious glacial till at depths of 4.6 to 24.4 meters. Lateral water movement occurs in a saturated soil zone that exists at the loess-till interface. Interception and/or sampling of water from both surface runoff and subsurface flow was done. In their study, watershed 2 (WS2) was conventionally tilled (33.5 ha) while watersheds 3 and 4 (WS3 and WS4) were contour-till planted (43.3 ha) and terrace-till planted (60.8 ha), respectively. About 65 head of cattle gleaned the cornstalks from WS3 and WS4 from mid-November to March each year. Figure 2 shows 10-year runoff-weighted concentrations of nitrate and ammonium for three time periods: April through June (fertilization, seedbed preparation, and crop establishment), July through November (the crop reproduction and maturation period), and December through March (the crop residue period), or periods 1, 2, and 3, respectively. U.S. water quality criteria in the NO₃^- and NH₄⁺ forms are currently 10.0 and 2.0 mg L^-1 respectively (Fletcher 1991, USEPA 1986). These values are shown on the dashed lines in Figure 2.

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Figure 2. Runoff-weighted concentrations of NO₃^--N and NH₄⁺-N in surface flow by seasonal period. Dashed lines represent current water quality standards. (Alberts and Spomer 1985)

Highest NO₃^- concentrations from the till-planted watersheds (WS3 and WS4) occurred during July through November (period 2), perhaps because evaporative drying was moving previously applied fertilizer salts to the soil surface. Preplant applications of fertilizer for the conventionally tilled watershed (WS2) had been incorporated with a disk. Ammonium N concentrations were generally higher during December to March (period 3). This probably resulted from cattle manure and leaching of NH₄⁺ from partially decomposed cornstalks.

Issues illustrated by this study include the need to place fertilizer below the soil surface while still maintaining residue cover for soil erosion control. Fall and winter livestock grazing of crop residues likely contributes to N loss in runoff since the manure and urine may be deposited on frozen ground.

Effect of Nitrogen Source on Fate and Transport

There are numerous N-containing compounds that can be used as N-fertilizers. However, when availability, economics, convenience, and effectiveness are considered, the compounds usually considered are those included in Table 1.

Table 1. Nitrogen fertilizer materials, their formulas and chemical analysis

Material	Chemical formula	Chemical analysis
Anhydrous ammonia	NH3	82
Ammonium nitrate	NH4NO3	33.5
Ammonium sulfate	NH4SO4	21
Diammonium phosphate	(NH4)2H2PO4	18-21
Monoammonium phosphate	NH4+H2PO4	11
Calcium nitrate	Ca(NO3)2	15
Calcium cyanamide	CaCN2	20-22
Potassium nitrate	KNO3	13
Sodium nitrate	NaNO3	16
Urea	CO(NH2)2	45
Urea-ammonium nitrate	CO(NH2)2+NH4NO3	28-32

Anhydrous ammonia, or gaseous NH₃, is a very important direct-application N-fertilizer. Gaseous NH₃, when in contact with moist soil, dissolves in and reacts with soil water to form NH₄⁺ and OH^- ions. The pH is increased dramatically immediately around the application zone of anhydrous NH₃. Therefore, depending upon the buffering capacity of the soil and the resulting soil pH, an equilibrium is approached between soil solution NH₄⁺ and NH₃ in the soil solution and gaseous NH₃, as was discussed above. The gaseous NH₃ can be lost by volatilization into the atmosphere. In addition, if anhydrous NH₃ is placed in dry soil or at too shallow a depth, the NH₃ is also subject to volatilization. However, the N that is in NH₄⁺ form is readily sorbed to the CEC of the soil.

All of the compounds shown in Table 1 are highly water soluble. For those with NH₄ as part in their chemical formula, the NH₄⁺ will sorb to the CEC of the soil. Therefore, the primary transport mechanism for NH₄⁺ ions is in association with eroding sediments since they are sorbed to the CEC of the soil.

Urea and calcium cyanamide (Table 1) are organic forms of N that, when applied to soil, are acted upon by enzymes to mineralize the N in them to NH₄⁺ ions. Once in NH₄⁺ form, the N in these two fertilizers is also sorbed to the CEC of the soil and is subject to the soil-erosion transport process described above. Also, the N in other organic materials such as manures and crop residues is also mineralized to NH₄⁺, again being subject to transport with eroding sediments.

For compounds in Table 1 that have NO₃ as part of their chemical formula, the NO₃^- does not sorb to the CEC of the soil. Therefore, the primary transport mechanism for NO₃^- ions is with percolating water by leaching or surface runoff (including return flow). Nitrate that is leached below the crop root zone often ends up as a pollutant in ground-water supplies. Nitrate can also be dissolved in surface runoff water or in return-flow water that returns to the surface to become part of the runoff.

Primary and Secondary Flows of Nitrogen

Primary and secondary flows of N are very much a part of the animal/plant N-cycling ecosystem as was discussed above. The following discussion is focused upon cropland and surrounding ecosystems, but also relates to a livestock system. Figure 3 from Duxbury et al. (1993) illustrates some of the flows of N following input of 100 kg of fertilizer N. The primary flows are shown by dashed lines. In this example, 50 of the 100 kg are harvested in the crop and 50 are lost by the combination of leaching (25 kg), surface runoff (5 kg), and gaseous loss (20 kg, primarily denitrification). If 10% of the gaseous N loss is N₂O, then 2 kg N₂O-N would be generated in the primary cycle. Another estimate of average N₂O-N emission is about 1 kg of N₂O-N per 100 kg of N in the applied fertilizer (CAST 1992).

Secondary flows shown by the solid lines include feeding of the 50 kg of harvested N to animals, which might generate about 45 kg of manure N. The manure is returned to cropland to create a secondary flow of the original fertilizer N. Part of this secondary flow of applied fertilizer N is again removed from the field by the harvested crop; through gaseous losses as NH₃, N₂O, NO_x, and as N₂ gas; in surface runoff; and by NO₃^- leaching. However, about half of the manure-N is volatilized as NH₃ prior to or during manure application. Volatilized NH₃ is aerially dispersed to be eventually returned to and cycled through both natural ecosystems and cropland (Duxbury et al. 1993). Estimates are that, over the course of about 50 years, more than 80% of the N applied to a field will eventually return to the atmosphere through denitrification (Cole et al. 1993). Generally, more than 95% of this N returns to the atmosphere as N₂ gas, but some unknown amount is released as N₂O.

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Figure 3. A simplified flow of fertilizer-N through the environment (Duxbury et al. 1993)

Concerns about Nitrogen's Impacts on the Environment

There are many local sources of N that contribute to water-quality problems. Typical point sources might include sites related to the disposal of human and animal wastes, industrial sites (for example those related to food processing), and sites where handling and accidental spills of nitrogenous materials may accumulate. However, the literature leaves no doubt that in farmed areas, agricultural activities comprise the bulk of nonpoint sources. Agricultural sources can be generalized as high-density animal operations where feed is transported into a watershed and the resulting manure-N must be spread at rates in excess of crop nutrient requirements, and row-crop agriculture where fertilizer N is used to supplement crop-N needs.

The primary concern about the impact of N on the environment is leaching of nitrate (NO₃^-) into ground water. This concern largely results from the potential for health effects that may result from humans and ruminant animals drinking contaminated ground water (Follett and Walker 1989). A number of reviews and papers have been published concerning NO₃^- in the environment and its possible health effects (Aldrich 1984, Brezonik 1978, Cast 1985, Duijvenbooden and Matthijsen 1987, Fraser and Chilvers 1981, Jaffe 1981, Keeney 1982, and Viets and Hageman 1971).

When NO₃^- accumulates in ground water and is ingested in high enough amounts, potential adverse health effects may occur. These health effects are reported to include methemoglobinemia, possibly cancer, and possibly other adverse effects. Public health standards for nitrate (NO₃^-) in public drinking water supplies in the USA have been set at 45 mg L^-1, or 10 mg N L^-1 in the NO₃^- -N form.

Methemoglobinemia (popularly called blue-baby syndrome) results when ingested NO₃^- is converted to the NO₂^- ion in the oral cavity and stomach and then absorbed from the gastrointestinal tract into the blood. Nitrite in the blood becomes involved in the oxidation of hemoglobin (Hb) to methemoglobin (metHb). The ferrous iron (Fe⁺²) in the heme group is oxidized to ferric iron (Fe⁺³) which NO₂^- firmly bonds, thus inhibiting transport of oxygen by the blood. Infants younger than three months are highly susceptible to gastric bacterial NO₃^- reduction because they have very little gastric acid production and low activity of the enzyme that reduces metHb back to Hb. With animals, NO₃^- toxicity is primarily a problem with ruminants in which bacterial reduction of NO₃^- to NO₂^- occurs in the rumen during the first stage of digestion and the NO₂^- is absorbed through the oral and gastrointestinal tract into the blood. The formation rate of metHb varies considerably among species.

An association between NO₃^- intake and gastric cancer is suggested by Fine (1982) based upon the correlation of stomach cancer mortality rates against previously published data on daily NO₃^- intake in different countries (r = 0.88). In addition to dietary NO₃^- levels, other factors involving biotransformation of NO₃^- may influence formation of gastric cancer. These include dietary salt intake, thiocyanate (smokers versus non-smokers), iodide intake, age, acidity of the gastrointestinal tract, and use of medication. In their review, Duijvenbooden and Matthijsen (1987) report that a majority of studies show no correlation or in some cases a negative one between NO₃^- intake and stomach cancer. Eating certain vegetables, even though high in NO₃^-, appears to be associated with lower risk of stomach cancer. Persons with chronic gastritis, especially the atrophic form, or those with other gastrointestinal problems are a special risk group. Also, persons with iron deficiency or those with pernicious anemia are predisposed to stomach cancer and also have a high rate of NO₃^- reduction to NO₂^-.

Experimental evidence does not show NO₃^- and NO₂^-, in and of themselves, to be carcinogenic and it is presently impossible to make a scientifically reliable estimate of the risk of human cancer posed by exposure to NO₃^- in drinking water. Nitrite can give rise to the formation of N-nitroso compounds by reaction with "nitrosatable compounds," including secondary and tertiary amines and amides, N-substituted ureas, guanidines, and urethanes. Sufficient toxicological data are available to indicate that humans are likely susceptible to the carcinogenicity of these compounds and that contact with them should be minimized (Brezonik 1978).

The Terrestrial Environment

Nitrate that moves below the crop root zone is totally soluble and can potentially leach into ground water. Ground water flows in permeable geologic formations called aquifers which are natural zones beneath the earth's surface that often yield economically important amounts of water. In a very simple system, water and dissolved NO₃^- percolate below the root zone and through the intermediate vadose zone to an aquifer. From there, these waters can recharge deeper aquifers or discharge to streams or water bodies. Aquifers are subdivided based upon geology. A meaningful division, from the perspective of ground-water quality, is between confined and unconfined aquifers. Confined aquifers are separated from the earth's surface by an aquiclude or aquitard. Unconfined aquifers are not separated from the earth's surface by a flow-impeding layer and are therefore in contact with the atmosphere through the unsaturated zone. Aquifer systems are often complex. To minimize the amount of NO₃^- that may enter ground water, it is necessary to understand the aquifer system and then to identify and apply improved N-management practices to the recharge area of the aquifer. Structure of the aquifer system and subsequent flow patterns affect NO₃^- dilution, transport, and removal.

Ground water can rejoin the ground surface downslope and adjacent to a perennial stream, often along a riparian zone similar to that shown in Figure 4. In a riparian zone, the water table moves progressively toward the land surface and the intermediate vadose zone is lost as the stream channel is approached. During storms or wet periods, the water table can rise rapidly to intersect the land surface at some distance from the stream; discharge of ground water to the soil surface results. The system can be dynamic, with water table levels, extent of the saturated zone, and flow directions changing substantially and rapidly with precipitation (Pionke and Lowrance 1989). As the ground water and its dissolved NO₃^- move into the more biologically and chemically active soil zones, the NO₃^- becomes available for uptake by riparian vegetation. Also, if oxygen levels become limited, activation of soil biological and chemical regimes result in denitrification.

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Figure 4. Schematic of the vadose zone, aquifers, and flow directions in a typical riparian zone in a humid climate (Pionke and Lowrance 1991)

Ground water--Nitrogen is the nutrient of most concern in the contamination of ground water, primarily resulting from NO₃^- leaching. Leaching of NH₄⁺ is generally not important since it is strongly adsorbed by soil, except for sands and soils having low retention (cation exchange) capacities. However, NO₃^- is readily leached deeper into the soil profile, below the bottom of the root zone, and may eventually leach into ground-water supplies.

Although elevated concentrations of NO₃^- are most often observed at shallow water-table depths, long-term increases in deeper wells are possible where deep aquifers are recharged by NO₃^--rich waters. Movement of NO₃^- with percolating water, through the unsaturated zone,can be very slow and time required for present-day inputs of NO₃^- to reach the ground water reservoir may be many years. Schuman et al. (1975) observed an average rate of NO₃^- movement through silt soils (loess) of about one meter per year for the first few years after application of 448 kg N ha^-1 to corn in Iowa. Where 168 kg N ha^-1 (the recommended N rate) was applied, N did not accumulate beneath the crop-root zone. Ground water flows from areas of high pressure toward areas of low pressure (hydraulic head). Generally movement is slow and there is little mixing of contaminated ground water with noncontaminated ground water that flows through the saturated zone, as contaminants tend to remain concentrated in zones. However, because of the slow rate of movement and lack of dilution, contamination may persist for decades and centuries, even if input sources of NO₃^- are decreased or eliminated. Unfortunately, reclamation is technically and economically impossible in most cases (Keeney 1986).

Many sites of excessive NO₃^- accumulation are recognized. Viets and Hageman (1971) conducted a comprehensive review of studies in the USA. Substantial accumulations of NO₃^- were found in deep profiles of irrigated Colorado soils, except where alfalfa was the crop (Stewart et al. 1967). Muir et al. (1973) conducted a study of factors influencing NO₃^- content of ground water in Nebraska. Their data indicated that the quality of Nebraska water was not being materially influenced by agricultural use of commercial fertilizers up to that time except on sites of intensively irrigated sandy soils and in valley positions with a shallow underlying water table.

There are numerous sources of N in the environment. Keeney (1986) identified intense land-use activities (e.g., irrigation farming of high value crops, high density of animal operations, or septic tank systems) as causes of excessive NO₃^- in ground water. Irrigation of cropland is widely practiced in the USA, particularly in the more arid West and in the Southeast where economic returns are high. The review by Pratt (1984) shows that in situations where roots have access to the entire soil solution, NO₃^- is not leached unless excess fertilizer N is added or the soils are over-irrigated.

Two general approaches to minimize NO₃^- leaching into ground water are: 1) optimum use of the crop's ability to compete with processes whereby plant-available N is lost from the soil-plant system, and 2) direct lowering of the rate and duration of the loss processes themselves. Key elements of the first approach are to assure vigorous crop growth and N-assimilation capacity, and to apply N in phase with crop demand. The second approach might include use of nitrification inhibitors or delayed release forms of N to directly lower potential leaching losses. In addition, realistic crop-yield goals must be selected. Olson (1985) emphasizes that a realistic yield goal would be no more that 10 percent above recent average yield for a given field or farm. Such a yield goal will still likely be difficult to achieve because of limitations imposed by environmental factors and/or the farmer's own operational skills.

Surface water--Agricultural production has been identified as a major source of nonpoint source pollution in U.S. lakes and rivers that do not meet water-quality goals. Sediment is the largest single type of pollutant followed by nutrients (NRC 1993). As has been discussed above, much of the N that enters lakes and rivers is associated with eroding sediments (NH₄⁺), eroding soil organic matter (organic forms of N and NH₄⁺), and surface runoff water (primarily dissolved NO₃^-). The water that runs over the soil surface during a rainfall or snowmelt event, by rill or sheet flow or even high-order channelized flow, may have a relatively high concentration of organic-N related to suspended particulate matter, but it is typically quite low in NO₃^- concentration.

The high NO₃^- flux that often occurs in streams draining agricultural land does not come from the overland runoff, but primarily from the ground-water contributions (including tile-drainage effluent) to streamflow. During discharge events the ground water and its NO3( load will include shallow interflow (sometimes referred to as "subsurface runoff"). However, most of the time, deeper baseflow that rejoins surface water provides the major contribution of NO₃^- (Hallberg 1989).

Stream water quality data from 904 nonpoint-source-type watersheds across the USA were summarized by Omernik (1977). The watersheds ranged in character from forested areas to urbanized regions, to areas dominated by row-crop agriculture. The data were correlated with land use and, as shown in Figure 5, especially the inorganic-N concentrations are directly related to the amount of the watershed used for agriculture. These data are over a decade old now; however, reviews of temporal trends show significant increases in NO₃^- since then (Hallberg 1989). Referring to Figure 5, concern about the long-term environmental impact may not only need to be the increasing loads of soluble N, but also the dramatic change in the proportion of the particulate and soluble-N concentrations. In forest and range systems the major N load was as organic-N, much of it in the particulate fraction (related to organic matter), but now the major load in agricultural areas is as soluble NO₃^-.

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Figure 5. Land use and mean inorganic and total nitrogen concentrations from stream data from 904 nonpoint-source-type watersheds (Omernik 1977)

The Aquatic Environment

Nitrogen can be transported into aquatic systems from airborne, surface, underground, and in situ sources (Table 2). Agricultural sources of N can arrive in surface water via airborne dust from wind erosion and through gaseous transport of NH₃ volatilized from livestock manure or from some fertilizer materials. Surface sources of N from agriculture are perhaps the best understood, and N delivered with eroded soil sediments is a major source.

Table 2. Sources and sinks for the nitrogen budgets of aquatic systems

Sources

Sinks

Airborne Rainwater Aerosols and dust Leaves and miscellaneous debris Surface Agricultural drainage including tile drainage Water erosion of sediment from agricultural land Animal waste runoff Marsh drainage Runoff and erosion from forest and rangeland Urban storm water runoff Domestic waste effluent Industrial waste effluent Wastes from boating activities Underground Natural ground water Subsurface agricultural and urban drainage Subsurface drainage from septic tanks In situ Nitrogen fixation Sediment leaching

Effluent loss Ground water recharge Fish harvest Weed harvest Insect emergence NH3 volatilization Evaporation (aerosol formation from surface foam) Denitrification Sediment deposition of detritus Sorption of ammonia onto sediments

Ground water delivery of NO₃^- to lakes and streams is no doubt very important, but difficult to gauge. In situ sources include biological N fixation, such as by blue-green algae, and the leaching of N from lake sediments. An additional source of N and other nutrients is from wild aquatic birds; however, their role in the nutrient regime of a water body may be more that of cycling agents than of direct sources.

Aquatic plants require a number of nutrients for growth, but N and phosphorus (P) appear to be the ones accounting for most of the excessive growth. However, under conditions where excessive growth is not occurring, nutrition of aquatic plants is necessary to provide sustenance for the aquatic food chain. The plants provide food for various aquatic invertebrate and vertebrate populations, which in turn support additional aquatic populations.

When waters become too enriched by nutrients, the aquatic environment can become eutrophic, with luxuriant growth or "bloom" of algae and macrophyte growth to levels that can choke navigable waterways, increase turbidity, and depress dissolved oxygen concentrations. Rapid growth of algae is the greatest and most widespread eutrophication problem. When a large mass of algae dies and begins to decay, the oxygen dissolved in water is depleted and certain toxins are produced; both of which can kill fish. The complexities ofeutrophication are that nutrient status of various species of algae can vary from lake to lake or even from different areas and depths of the same lake on the same day. Excess algal growth can create obnoxious conditions in ponded waters, increase water treatment costs by clogging screens and requiring more chemicals, and cause serious taste and odor problems.

Sawyer (1947) was the first to propose quantitative quidelines for lakes. He suggested that 0.3 mg L^-1 of inorganic N and 0.015 mg L^-1 of inorganic P are critical levels above which algal blooms can normally be expected in lakes. However, the EPA has not developed nutrient criteria or recommended methodologies for protecting waterbodies from excessive nutrient loading. National criteria that are available for NO₃^-, NO₂^-, and NH₃ were established to protect human health and aquatic life from toxic effects associated with these compounds. They do not address eutrophication or impairments to recreational uses; such as swimming, fishing, and boating (Tetra Tech, Inc. 1994).

Under natural conditions, NO₃^- and NO₂^- occur in moderate concentrations and have little toxicological significance for aquatic life. Because the levels that are toxic to aquatic life are much higher than those expected to occur naturally in surface waters, restrictive water quality criteria for these elements have not been recommended. As was discussed above in the section "Concerns about Nitrogen's Impacts on the Environment," the primary water-quality concern about these two forms of N stems from their potential health effects on humans and ruminant animals associated with the drinking of contaminated water.

On the other hand, NH₃ is highly toxic to aquatic organisms. Acute toxicity in fish causes loss of equilibrium, hyperexcitability, increased breathing, cardiac output, convulsions, coma, and death, if concentrations are extreme. Chronic toxic effects include reduced hatching success and growth rates, and developmental or pathological changes in gill, liver, and kidney tissues (USEPA 1986).

Global Climate

Nitrogen is the one nutrient that may potentially affect global climate directly when it is emitted to the atmosphere in the form of N₂O (nitrous oxide). Under favorable environmental conditions, Nitrosomonas spp. bacteria in the soil readily transform NH₄⁺ to NO₂^- that in turn is transformed by Nitrobacter spp. bacteria to NO₃^- (Figure 1). The small quantity of N₂O produced during nitrification of NH₄⁺ in aerobic soils is a direct metabolic product of chemoautotrophic NH₄⁺-oxidizing bacteria or results from other soil processes dependent upon these organisms as a source of NO₂^- (Tortoso and Hutchinson 1990). Many groups of organisms are capable of production and consumption of N₂O. Biogenic production in soil is the principal source of atmospheric N₂O. Anaerobic soil processes, rather than nitrification (an aerobic process), are the principal biogenic sources of atmospheric N₂O (Freney et al. 1979; Goodroad and Keeney 1984; Klemedtsson et al. 1988).

Denitrification is a bacterial process, during which NO₃^- or NO₂^- is reduced to gaseous N species NO, N₂O or N₂, and which is capable of producing and consuming N₂O and NO (equation 1).

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The general conditions required for denitrification to occur include: (a) presence of bacteria possessing the metabolic capacity; (b) availability of suitable reductants such as organic C; (c) restriction of O₂ availability; (d) availability of N oxides, NO₃^-, NO, or N₂O (Firestone and Davidson 1989). The strongly bonded N₂ gaseous molecule has an extremely large reservoir and is the predominant gas (78 percent by volume) in the atmosphere, whereas the atmospheric reservoir for N₂O is about 300 ppb (Smith et al. 1990). Nitrogen fertilizer additions, in either the NH₄⁺ or NO₃^- form, potentially contribute to release of N₂O to the atmosphere, especially where excess NO₃^- accumulates in the soil profile and is available for denitrification.

Because N₂O is the greenhouse gas of concern, the proportion of N₂O produced relative to N₂ under denitrifying conditions becomes of special concern. A number of factors affect the proportion of N₂O to N₂. A model by Betlach and Tiedje (1981) predicts accumulation of N₂O whenever one of the factors shown in Table 3 slows the rate of overall reduction.

Table 3. Factors affecting the proportion of N₂O and N₂ produced during denitrification (CAST 1992)

Factor	Will increase ratio of N2O to N2
[NO3-] or [NO2-]	Increasing oxidant
[O2]	Increasing O2
Carbon	Decreasing C availability
pH	Decreasing pH
[H2S]	Increasing sulfide
Temperature	Decreasing temperature
Enzyme status	Low N2O reductase activity

Nitrogen Use Efficiency

Fertilizing crops for crop-N uptake that will be near the point of maximum yield generally is an economically and environmentally acceptable practice. The objective is to lower the rate and duration of the loss processes themselves. Practices that decrease the opportunity for loss processes to occur will help decrease the amount of N that may be lost to the environment. In some cases improved efficiency is achieved by using less nutrients and in other cases it can be achieved by increasing the yield while using the same amount of N-input. In either case, the goal is to decrease the total residual mass of N in the soil. A second approach is to keep the residual N in the soil-crop system by curtailing the transport processes (leaching, runoff, erosion, and gaseous losses) that carry pollutants out of the soil-crop system or by increasing the mass of inputs immobilized or degraded in the soil crop system.

Current practices generally involve supplying crop-N needs in one, two, or three fertilizer applications. Most N is now applied either in the Nh3+ form or as urea, which normally hydrolyzes to NH₄⁺ within a matter of hours to days. In NH₄⁺ form the N is sorbed to the soil particles and cannot be leached from the soil. However, micro-organisms convert NH₄⁺ to NO₃^- which does move with water. This conversion may occur within a period of two or more weeks depending upon temperature and moisture conditions. If uptake by a crop has not removed NO₃^- from the soil solution, its concentration can become quite high. Rainfall or irrigation during such periods tends to displace this solution below the root zone.

Lack of synchronization of N supply (from either fertilizer or organic sources) with N needs and uptake patterns of crops is a primary cause of increased N leaching below the crop root zone. Other major causes are over-applications of N-fertilizer because of miscalculations of the amount needed or to insure against N-losses and/or under-fertilization.

Producers try to maximize their profits by applying inputs at economically optimum rates. Such optimum rates of application are closely related to rates that are optimal for crop growth, but are not necessarily the same. Economically optimum rates can be greater than optimum rates for crop growth because of uncertainties about outcomes and prices of inputs and the crop.

Producers are often thought to apply N at rates greater than those required for optimal crop growth as insurance against making a wrong decision that may lead to lower yields. Figure 6 shows average economic losses caused by under application of N and the gains from the overapplication of N as insurance. Bock and Hergert (1991) concluded that economic incentives for N application at the optimum rate are not great, particularly when the yield response to N is highly variable and N-price/crop-price ratios are low.

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Figure 6. Economic return from insurance nitrogen and deficit nitrogen applications (Bock and Hergert 1991)

Fertilization and Cropping System Strategies

Crop nitrogen requirements ( Crop N uptake determinations provide the basis for estimating N requirements which can then be adjusted to compensate for other N sources. The crop N requirement factor can be considered as crop N uptake near maximum yield. Thus, on a kg N uptake per 1000 kg grain basis, typical values for corn and wheat are near 25 and 40 respectively. Yield goals, when possible, should be based on prior history of the soil and level of farm management. A realistic yield goal is that obtained by taking average production over a 5-year period and adding no more than 5 percent to that production.

Improved N-use efficiency will require synchronization of soil N availability and crop N requirement. Thus, cultural practices, soil-water status, crop pests, and many other factors that affect crop N-uptake patterns will complicate management decisions. Genetic selection for improved N efficiency in crops such as corn and sorghum may reduce N requirements, but efforts to quantify N_x genotype interactions can easily be dominated by environmental interactions.

Accounting for all nitrogen sources--Nitrogen budgets provide a valuable framework to quantify and examine N inputs and losses for agricultural production systems (see also Figure 1). Accounting for the major sources of N to cropping systems is especially important. The following are some of the sources that should be considered. Fertilizer N inputs and amounts are easily determined and managed; a number of sources are listed in Table 1 above. Organic wastes are an important N source. Organic wastes available for use on cropland in the United States include livestock wastes, crop residues, sewage sludge and septage, food processing wastes, industrial organic wastes, logging and wood manufacturing wastes, and municipal refuse. Animal manures and crop residues account for the majority of organic wastes applied to cropland. Uncertainties about manure N inputs arise not only because the N content is related to livestock type, age and health, but also because, once excreted, the N content can change considerably depending on type and amount of bedding, type and time of manure storage, and manure management and placement when being applied. The best way to overcome these uncertainties is through the use of manure analysis and calibration of application equipment. Manure credits are often used to account for N that becomes available from applied manure. Biological N-fixation (BNF), especially by legumes, can be an especially important source of N. Although the importance of BNF has been known for centuries, there are few quantitative methods for estimation of BNF. Currently, the method most used is that of recognizing BNF by legumes with legume credits. Nitrate contained in irrigation water is available to the crop and should be considered when making fertilizer recommendations. Crop utilization of NO₃^- from irrigation water is greatest when the crop N requirement is the greatest and when other sources of N in soil are not excessive. Atmospheric additions are another source of N to crops. The mechanism of additions that are identified include N dissolved in precipitation, dry deposition, and direct plant absorption of gaseous NH₃. Contributions of residual soil N require soil testing for NO₃^- and NH₄⁺ within the root zone and will be discussed below. Nitrogen mineralization is the term given to biological decomposition of organic material in soils and their conversion to inorganic forms. The contribution of organic forms is significant and again will be discussed below.

Soil-nitrogen availability tests--Available soil N represents residual N in the soil profile, plus N mineralized from the soil organic matter during the growing season. While residual N has proven to be a useful index in certain regions of the USA, no generally accepted index exists for N mineralization. Obviously, such development would represent a major advance for avoidance of excessive fertilizer N applications. A complement to a soil-N test may be a plant-tissue-N test. An attractive feature of tissue tests is that the plant root system tends to integrate spatial variability of soil N supplying power over a relatively large field volume.

Soil organic-nitrogen availability--A significant portion of the crop's N-need is supplied by mineralization of soil organic matter during the growing season. Various N-availability indexes exist, but they typically provide qualitative rather than quantitative measures of soil organic-N availability. Early concepts of an N-availability index have been modified in a number of ways; but, to date, no soil organic-N availability procedure has received general acceptance from a soil test standpoint. Ultimately, a systems-type, mass-balance N approach may be the best alternative. The present recommendation is to follow pertinent N-fertilizer guides which have been developed locally for specific crop needs and soil areas.

Nitrification inhibitors--The NH₄⁺ ion is sorbed to the CEC of the soil, whereas the NO₃^- ion is not, and can be readily leached or denitrified. Both forms can be readily utilized by crops. Nitrification inhibitors include chemicals added to soils to stabilize fertilizer applied as NH₃ or in the NH₄⁺ form by inhibiting the activity of the Nitrosomonas bacteria in the first step of the nitrification process.

Control/slow release fertilizer--Methods of altering the release of N from soluble materials has been to coat water-soluble N fertilizer with less water-soluble materials in order to retard entry of water into the particle and the movement of N out. Coatings applied to soluble N materials generally have been of three types: 1) Impermeable coatings with small pores that allow slow entrance of water and slow passage of solubilized N out of the encapsulated area; 2) Impermeable coatings that require breakage by physical, chemical, or biological action before the N is dissolved; and 3) Semipermeable coatings through which water diffuses, creating internal pressures sufficient to disrupt the coatings. Sulfur-coated urea (SCU) has been developed in recent years as a product with the characteristics of slow N release. Elemental sulfur (S) was chosen because of its relatively low cost and ease of handling. Newer control-release N-fertilizer materials are also being developed and marketed (Shaji and Gandeza 1992). These newer materials have polyolefin resin coatings. The coatings can be tailored to provide a range of N-release rates that are suitable for a variety of cropping systems. However, further field research is needed to ensure the utility of these newer materials for U.S. cropping systems.

Conservation tillage--Use of conservation or reduced-tillage (including no-till) continues to increase as an alternative for nearly all forms of crop production. Management systems which maintain crop residues at or near the soil surface have several attractive features, including less on-farm energy use, more available soil water, and reduced soil erosion. However, adoption of conservation tillage practices may result in some N moving from the soil-plant system into the environment under certain conditions.

There is no question that conservation tillage is effective in decreasing particulate N losses associated with soil erosion and surface-water runoff as has been discussed above. However, effects of conservation tillage on leachable N are not so well delineated as are surface losses. Generally, conservation tillage provides a wetter, cooler, more acidic, less oxidative soil environment. Under such conditions, processes of ammonification and denitrification may be favored over nitrification. Conversely, for NO₃^- that is already present, the leaching potential may be greater under conservation tillage. This is because more undisturbed soil- macropores exist for NO₃^- and water movement. Increased water flow into and through the root-zone has been observed under no-till compared to conventionally tilled soils. This higher flow has been attributed to decreased water evaporation because of surface residues and increased numbers of undisturbed channels (e.g. earthworm and old roots) continuous to the soil surface. The surface mulch enhances the environment for earthworms and the lack of tillage preserves existing channels for several years.

Rotations, cover crops, and nitrogen-scavenging crops--Rotations and cover crops, historically used as a means of conserving soil or of providing an organic N source or both, have received renewed interest as an aid in avoiding excessive N losses to the environment. Whereas monocultures of grain crops (e.g. corn and wheat) require high inputs of fertilizer N, such inputs can be decreased with crop rotations which require less or fix atmospheric N. Because less excess profile N may be expected with a rotation, there should be less potential for N leaching. An exception may be under certain rotation-fallow conditions designed to conserve water in drier areas.

Winter cover crops can be effective in absorbing both NO₃^- and available water during the fall, winter, and spring, thereby decreasing the N leaching potential. When the cover crop is returned to the soil, some of the absorbed N is then available to the following crop. Both legumes and nonlegumes are used, but from a strictly N leaching standpoint. While an annual crop, such as rye, can be effective in scavenging excess available N from within crop rooting zones, deep-rooted perennials should be considered for NO₃^- accumulation below normal rooting depths. Alfalfa, with a potential rooting depth in excess of 15 feet, is a crop which merits particular attention.

Filter strips--Vegetative filter strips, buffer strips and riparian zones all remove sediment, organic matter, and other pollutants from runoff and waste waters. Under field conditions, excess runoff from terraces is frequently diverted to a grassed waterway. Such a waterway on an unterraced field may follow the natural lie of the land, but at a terrace outlet its slope and direction are commonly designed and mechanically shaped, like the terraces, to reduce the flow velocity and therefore the transport capacity of the runoff. The sediment and its associated pollutants are then removed from the runoff by filtration, deposition, infiltration, sorption, decomposition, and volatilization processes.

The effectiveness of filter strips in removing sediment and particulate N is well established. Less certain is the effectiveness of filter strips for removing soluble N in runoff. Filter strip vegetation assimilates mineral N transported by runoff water most effectively during times of active growth, but less during other times of the year. Also, some denitrification may be occurring. In riparian areas, however, plants start into growth earlier and remain in active growth longer than on uplands, and the greatest efficiency is generally attained by a combination of woody species, forbs and grasses suited to the site or region. Scavenging of N from underground water and the vertical horizon by riparian vegetation, especially by deeper rooted plants, may be important for removing dissolved N in surface and subsurface flows before the N is transported into streams and lakes.

Source Areas and In-Field Targeting

Water-quality impact zones for N are wells, ground-water supplies, streams, and surface water bodies. Because 95 percent of rural inhabitants and substantial livestock populations consume ground water, NO₃^- concentration is most important. Factors which control NO₃^- concentration, such as dilution and well position relative to the primary source areas, can greatly affect their impact on ground-water quality. In contrast, streamflow tends to mix ground-water discharge and surface runoff from different land uses and time periods, thus producing generally much lower and more stable NO₃^- concentrations.

Because the subsurface system is generally large and nonuniform in structure, function, or efficiency, it is much easier to focus on source areas rather than on the whole system. The source area is a bounded area or volume within which one process or a set of related processes dominate to provide excessive production (source), permanent removal (sink), detention (storage), or dilution of NO₃^-. Source area effects, by definition, are disproportionately large relative to the area or volume occupied. If the source area(s) can be identified and positioned relative to the generalized flow pattern within the system, there is a possible basis for estimating effects on an impact zone.

Systematic data on production practices, input use, and management systems are insufficient to do many of the assessments that are needed. However, the quantity and quality of soil survey data, climate data, and assessments of NO₃^- concentrations in various aquifers are increasing. Statistical techniques and simulation models used in conjunction with Geographical Information Systems (GIS) technology show promise in identifying and assessing NO₃^- leaching across regions (Christy 1992, Wylie et al. 1994). Models such as the Nitrate Leaching and Economic Analysis Package (NLEAP) (Shaffer et al. 1991) use farm management, soil, and climate information to estimate NO₃^- leaching at a farm or even the soil series level, thus allowing determination of potential NO₃^--leaching hot spots on the landscape when sufficient information is available. As technology continues to improve, it should become possible to target improved practices to those areas, farm enterprises, fields within a farm, or even locations (hot spots) within a field that cause the most damage, and thus to substantially reduce losses of N to the environment.

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