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RCA III
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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 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, NOx, 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)
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).
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)
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.
Sources | Sinks |
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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).
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.
Factor | Will increase ratio of N2O to N2 |
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[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 |
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)
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 Nx 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.
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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