Natural Resources Conservation Service

Fate and Transport of Nutrients: Phosphorus

Working Paper No. 8

Andrew Sharpley
USDA, Agricultural Research Service
National Agricultural Water Quality Laboratory
Durant, Oklahoma

October 1995

Contents

Soil Phosphorus Cycle
Transport Processes
Effect of Phosphorus Sources on Fate and Transport
Impacts of Phosphorus on the Environment
Use-efficiency to Decrease Negative Impacts
Conclusions
References

Phosphorus (P) is an essential element for plant growth. However, immobilization of soil P in inorganic and organic forms unavailable for crop uptake necessitates P amendments as fertilizer or animal manure to achieve desired crop yield goals. Although P is not directly toxic, the continued application of P to agricultural land and its subsequent movement to surface waters in runoff can accelerate eutrophication. This can impair water use for industry, recreation, drinking, and fisheries, due to the increased growth of undesirable algae and aquatic weeds. Although nitrogen (N) and carbon (C) are also associated with accelerated eutrophication, most attention has focused on P, due to the difficulty in controlling the exchange of N and C between the atmosphere and a water body, and fixation of atmospheric N by some blue-green algae. Thus, P is often the limiting element and its control is of prime importance in reducing the accelerated eutrophication of surface waters.

In areas of intensive crop and livestock production, continual P applications as mineral fertilizer and manure have been made at levels exceeding crop uptake (Sharpley et al., 1994b). As a result, surface soil accumulations of P have occurred to such an extent that the loss of P in surface runoff has become a priority management concern. In several cases, the capacity of the soil to adsorb further P additions has become limited, with increased losses of P in ground water observed in Florida (Federico et al., 1981) and the Netherlands (Breeuwsma and Schoumans, 1987).

Before we can develop agronomically and environmentally sound agricultural systems for P, we need to understand what forms of P occur in soils, the dynamics of cycling between forms of differing bioavailability (i.e., available for uptake by plants and aquatic biota), and the processes controlling soil P removal by and transport in runoff. Using this information, we can assess how to identify and manage agricultural P to maximize soil productivity, while minimizing environmental impacts.

Soil Phosphorus Cycle

Forms

Soil P exists in inorganic and organic forms (figure 1). These forms are characterized by chemical extractions and relative lability assigned as to the chemical species extracted. Such fractionation of soil P is based on the premise that extractants of increasing acidity and alkalinity sequentially remove P of decreasing lability or bioavailability (Hedley et al., 1982). Inorganic P forms are dominated by hydrous sesquioxides, amorphous and crystalline Al and Fe compounds in acidic, noncalcareous soils and by Ca compounds in alkaline, calcareous soils (figure 1). Organic P forms include relatively labile phospholipids, inositols and fulvic acids, while more resistant forms are comprised of humic acids (figure 1). However, the forms generalized in figure 1 are not discrete entities, as intergrades and dynamic transformations between forms occur continuously to maintain equilibrium conditions.

In most soils, the P content of surface horizons is greater than that of subsoil due to the adsorption of added P and greater biological activity and accumulation of organic material in surface layers. However, soil P content varies with parent material, texture, and management factors, such as the rate and type of P applied and soil cultivation. These factors also influence the relative amounts of inorganic and organic P. In most soils 50 to 75% of the P is inorganic, although this fraction can vary from 10 to 90%.

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Figure 1. The soil P cycle: Its components and measurable fractions
(adapted from Stewart and Sharpley, 1987).

Transformations

I. Dryland Soils

With the application of P, available soil P content increases as a function of certain physical and chemical soil properties. The portion of P remaining as plant available P (resin P) 6 months after application decreased as clay, organic C, Fe, Al, and CaCO₃ content increased for over 200 widely differing soils (table 1: Sharpley et al., 1984; 1989). With an increase in degree of soil weathering, represented by soil taxonomic and other related properties, a general decrease in availability of applied P was evident.

TABLE 1 Phosphorus fertilizer available (as resin phosphorus) six months after application

Related properties	Number of soils	Availability
		Mean	Range
		Percent
Calcareous CaCO3	56	45	11-72
Slightly weathered Base saturation Available P pH	80	47	7-74
Moderately weathered Clay Available P Organic C	27	32	6-51
Highly weathered Clay Extractable Al Extractable Fe	40	27	14-54
Source of data: Sharpley, 1991; Sharpley et al., 1984a, 1989

Even though inorganic P has generally been considered the major source of plant available P in soils, the incorporation of fertilizer P into soil organic P (McLaughlin et al., 1988) and lack of crop response to fertilizer P due to organic P mineralization (Doerge and Gardner, 1978) emphasize the importance of organic P in soil P cycling (figure 1). Tate et al. (1991) also found that labile organic P mineralization was an important source of P to pasture in both low-P and high-P fertility soils in New Zealand. Amounts of organic P mineralized in temperate dryland soils range from 5 to 20 kg P ha^-1 yr^-1 (Stewart and Sharpley, 1987). Mineralization of soil organic P tends to be higher in the tropics (67 to 157 kg P ha^-1 yr^-1), where distinct wet and dry seasons and higher soil temperatures enhance microbial activity.

The role of microbial biomass P as a dynamic intermediary between organic and inorganic forms is evident from figure 1. With the development of fumigation-extraction techniques to measure soil microbial biomass P (Brookes et al., 1982; Hedley and Stewart, 1982), its importance in P cycling has been quantified (McLaughlin et al., 1988; Stewart and Tiessen, 1987). In a study of P cycling through soil microbial biomass in England, Brookes et al. (1984) measured annual P fluxes of 5 and 23 kg P ha^-1 yr^-1 in soils under continuous wheat and permanent grass, respectively. Although biomass P flux under continuous wheat was less than P uptake by the crop (20 kg P ha^-1 yr^-1), annual P flux in the grassland soils was much greater than P uptake by the grass (12 kg P ha^-1 yr^-1). Clearly, microbial P plays an important intermediary role in the short-term dynamics of organic P transformations, which suggests that management practices that maximize the buildup of organic matter during autumn and winter may reduce external P requirements for plant growth during the following spring and early summer.

II. Wetland Soils

The rate and extent of inorganic and organic P transformations in wetlands are modified by intermittent aerobic and anaerobic conditions, compared to the dryland soils described above. Under aerobic conditions the solubility of P associated with amorphous and sesquioxide Al and Fe compounds can be increased, but some P associated with crystalline Al and Fe oxides is desorbed only under extended waterlogged conditions (figure 2). Although Al-P complexes are not affected by oxidation-reduction reactions brought about by aerobic-anaerobic conditions, pH and organic matter influence the solubility of these P forms. Thus, Fe speciation tends to dominate the dynamics of P solubility in many wetland soils. Ferric oxyhydroxides in the surface oxidized zone of a waterlogged soil can act as sink for P in both the overlying water column and underlying anaerobic soil, as represented in figure 2. Thus, the thickness of this surface layer, which is dependent on oxygen-demanding species at the interface, will to a large extent determine the mobility of P associated with Fe complexes in wetland soils.

Organic P mineralization may be enhanced by alternate soil wetting and drying cycles, changes in soil pH, and an increase in microbial activity. Reddy (1983) estimated that organic P mineralization amounted to about 38 to 185 kg P ha^-1 yr^-1 for organic soils in central Florida and 16 to 23 kg P ha^-1 yr^-1 in south Florida. As a result, flooding these organic soils increased P release into drainage water by about 4 to 8 times that from drained soils.

 

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Figure 2. Generalized scheme of P transformation in a wetland system.

In general, the bioavailability and mobility of P in wetland soils under aerobic conditions are greater than for aerobic or dryland soils. This enhances the potential for P movement in runoff and drainage water from wetland soils. Wetland soils can function as sinks and sources of P depending on water residence time, sediment physiochemical properties and vegetative assimilation (Reddy et al., 1994; Richardson, 1985). When the dissolved P concentration of water flowing into wetlands is greater than that present in the pore water of wetland soils, P is retained by Al, Fe, organic matter, and to a lesser extent by Ca complexes. However, with low P loadings, wetland soils can act as a P source, releasing P to the water column (Bostrom et al., 1982; Khalid et al., 1977).

Plant and Animal Uptake

Several environmental factors affect plant uptake of P from any source, soil or fertilizer (Munson and Murphy, 1986). These include temperature, soil compaction, soil moisture, soil aeration, soil pH, type and amount of clay content, P status of soil, and status of other nutrients in soil. When soil temperatures are low during early plant growth, P uptake is reduced. Soil compaction reduces pore space and consequently water and oxygen, which in turn reduces P uptake. Soil pH greatly affects the availability of P to plants, with P being tied up by Ca at high pH and by Fe and Al at low pH. Soils with high clay content tend to fix more P than sandy soils with a low clay content. Thus, more P needs to be added to raise the soil test level of clay soils than loam and sandy soils. In addition, the presence of ammonium enhances P uptake by creating an acid environment around the root when ammonium ions are absorbed. High concentrations of ammonium-N in the soil with fertilizer P may interfere with and delay normal P fixation reactions, prolonging the availability of fertilizer P (Murphy, 1988). Thus, many factors can affect P availability to crops.

Uptake of P by animals from feed materials can influence the fate and transport of P in soils. A generalized P balance and efficiency of plant and animal uptake of P for the United States and several European countries indicates the potential for P accumulation in agricultural systems (table 2). Although the magnitude of P input and output varies between countries, the relative efficiency of plant and animal removal is similar (table 2). In spite of the relatively efficient P utilization of 56 to 76% in crop production, P utilization by total agriculture is only 11 to 38% (table 2). In contrast to crop production, P efficiency with animal production is only 10 to 34%. Thus, the efficiency of total agriculture is dominated by that of animal production, as 76 to 94% of the total crop production is fed to animals (in addition to P additives) with low P utilization efficiencies. Animal-specific P excretion rates substantiate this at 70 to 80% for dairy cows and feeder pigs and 87% for poultry (Isermann, 1990). Clearly, agricultural systems that include confined animal operations can determine the overall efficiency of P utilization in agriculture and thereby the magnitude of P surpluses or potential soil accumulations.

Table 2. Phosphorus balance and efficiency of plant and animal uptake of phosphorus for selected countries

	Input	Output	Surplus	Efficiency of
				Plant uptake	Animal uptake	Total uptake
	kg P ha-1 yr-1			Percent
U.S.	39	13	26	56	15	33
Netherlands	143	55	88	69	24	38
E. Germany	79	8	71	59	10	11
W. Germany	84	29	55	76	34	35
Source of data: for U.S--NRC, 1993; for European countries--Isermann, 1990

Transport Processes

Forms

The transport of P in runoff can occur in dissolved (DP) and particulate (PP) forms (figure 3). Particulate P encompasses all phase forms, which include P sorbed by soil particles and organic matter eroded during runoff and constitute the major proportion of P transported from cultivated land (75 to 90%). Runoff from grass or forest land carries little sediment, and is, therefore, generally dominated by the dissolved form. While DP is, for the most part, immediately available for biological uptake (Nurnberg and Peters, 1984; Walton and Lee, 1972), PP can provide a long-term source of P for aquatic biota (Carignan and Kalff, 1980; Wildung et al., 1974). Together DP and bioavailable PP (BPP) constitute the bioavailable P (BAP; P available for uptake by aquatic biota) content of runoff. Bioavailable P can be estimated using iron oxide-impregnated filter paper (Fe-oxide strip) as a P-sink to adsorb BAP from a sample of runoff or sediment (Sharpley, 1993).

Runoff and Erosion

The transport of DP in runoff is initiated by the desorption, dissolution, and extraction of P from soil and plant material (figure 3). These processes occur as a portion of rainfall interacts with a thin layer of surface soil (1 to 5 cm) before leaving the field as runoff (Sharpley 1985a). Although this depth is difficult to quantify in the field, it is expected to be highly dynamic due to variations in rainfall intensity, soil tilth, and vegetative cover. Several studies have reported that the loss of DP in runoff is dependent on the soil P content of surface soil. For example, a highly significant linear relationship was obtained between the DP concentration of runoff and soil P content (Mehlich 3) of surface soil (5 cm) from cropped and grassed watersheds in Arkansas and Oklahoma (figure 4). A similar dependence of the DP concentration of runoff on Bray-1 P was found by Romkens and Nelson (1974) for a Russell silt loam in Illinois (r² = 0.81) and on water-extractable soil P (r² = 0.61) of 17 Mississippi watersheds by Schreiber (1988) and 11 Oklahoma watersheds by Olness et al. (1975) (r² = 0.88).

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Figure 3. Processes involved in the transfer of P from terrestrial to aquatic ecosystems.

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Figure 4. Effect of soil test P (Mehlich 3) on the dissolved-P concentration of runoff from several Arkansas and Oklahoma watersheds.

As the sources of PP in streams include eroding surface soil, streambanks, and channel beds, processes determining soil erosion also control PP transport (figure 3). In general, the P content and reactivity of eroded particulate material is greater than that of source soil, due to preferential transport of clay-sized material (<2 microns). Sharpley (1985b) observed that the plant-available P content of runoff sediment was on average 3-fold greater than that of source soil and 1.5-fold greater for total, inorganic, and organic forms.

Leaching

The P content of water percolating through the soil profile is genarally low due to sorption of P by P-deficient subsoils. Exceptions may occur in organic or peaty soils, where organic matter may accelerate the downward movement of P together with organic acids and Fe and Al (Dux-bury and Peverly, 1978; Fox and Kamprath, 1971; Miller, 1979). Similarly, P is more susceptible to movement through sandy soils with low P-sorption capacities and in soils which have become waterlogged, where conversion of Fe (III) to Fe (II) content and organic P mineralization occur (Gotoh and Patrick, 1974; Ozanne et al., 1961).

An interaction of these biogeochemical processes contributes to inefficient P retention by several sandy Haplaquods in areas of Okeechobee Basin, Florida, with a high density of dairy farms (Graetz and Nair, 1994). A low P-sorption capacity of the surface soil and a lack of percolation of DP into high P-sorbing subsoils, due to high water tables, contribute to high DP concentrations in drainage discharged from these basins (Campbell et al., 1994; Reddy et al., 1994).

Because of the variable path and time of water flow through a soil with subsurface drain-age, factors controlling DP in subsurface waters are more complex than for surface runoff. However, soil P content has been shown to influence the loss of P in drainage water as well as surface runoff. For example, Sharpley et al. (1977) found that the amount of P extracted by 0.1 M NaCl from soil at the tile drain depth (40 to 50 cm) was related (r2 = 0.86) to the DP loss in tile drainage during storm events (figure 5). A similar dependence of DP concentration in tile drainage on the P sorption-desorption properties of subsoil material was found for Histosols in Florida (Hortenstine and Forbes, 1972), New York (Cogger and Duxbury, 1984), and Ontario (Nicholls and MacCrimmon, 1974), and for Haplaquolls in Ontario and Michigan (Culley et al., 1983).

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Figure 5. Relationship between the extractable soil P content (0.1 M NaCl) of subsoil (40-50 cm) and dissolved P loss in the drain discharge events from a grassland watershed in New Zealand.

Changes in Bioavailability during Transport

Transformations between DP and PP, occurring during transport in streamflow, can alter both the amount and bioavailability of P entering a lake, compared to edge-of-field losses (figure 3). These transformations are accentuated by the selective transport of fine materials, which have a greater capacity to sorb or desorb P and will thus be important in determining the bioavailability of P transported. In addition, P may be taken up by aquatic biota and PP deposited or eroded from the streambed with a change in streamflow (Meyer, 1979; Vincent and Downes, 1980). The direction and extent of P exchange between DP and PP will depend on their relative concentration in streamflow, sediments contacted, and rate of streamflow.

Once sediment settles to the bottom of a lake, however, sediment P bioavailability will be increased by development of reducing conditions at the sediment-water interface (Nurnberg et al., 1986; Syers et al., 1973). For example, in a study of the P dynamics of two shallow hypereutrophic lakes in Indiana, Theis and McCabe (1978) found that the DP concentration of lake water was reduced by sorption during aerobic periods and increased by release of sediment P during anaerobic periods. It is apparent, therefore, that changes in P bioavailability can occur between the point where it leaves a field and that where it enters a water body. Consequently, the extent to which transformations between DP and PP occur during streamflow must be considered in assessing the impact of P transported in runoff as a function of agricultural management on the potential biological productivity of a receiving lake.

Effect of Phosphorus Sources on Fate and Transport

Fate in Soil

Adsorption of P by soil occurs rapidly, and because of the high binding energy between soil and P, adsorption tends to dominate desorption. Thus, a general decrease in soil P availability occurs after P is applied (figure 6). If soil test P decreases below a critical level, desorption of unavailable P can occur, but usually at a rate too slow to satisfy crop P requirements. The critical soil solution P level of a given soil is determined by the content and activity of Fe, Al, and Ca compounds adsorbing P.

The continual long-term application of P in fertilizer and manure at levels exceeding crop requirements can raise soil test P above levels required for optimum crop yields (figure 7).

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Figure 6. Plant availability of phosphorus decreases as time after application lengthens.

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Figure 7. Relationship between crop yield, soil test P, and potential for environmental problems.

Once soil test P levels become excessive, the potential for P loss in runoff and drainage water is greater than any agronomic benefits of further P applications. In recent years, the number of soils with soil test P exceeding the levels required for optimum crop yields has increased in areas of intensive agricultural and livestock production. In 1989, several state soil test laboratories in the eastern United States reported that the majority of soils analyzed had soil test P levels in the high or excessive categories (figure 8). Although these categories vary between states, soil test P limits range from >10 to >75 mg P kg ^-1 for high and from >25 to >150 mg P kg ^-1 for very high.

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Figure 8. Percentage of soil samples testing high or above for P in 1989. Shaded states have 50% or more of soil samples testing in the high or above range (adapted from Potash and Phosphate Institute, 1990, and Sims, 1993a).

Of particular concern is the efficient utilization in locally limited land areas of the manure produced in confined animal operations. For example, Graetz and Nair (1994) found soil test P levels (double acid) of 453 mg P kg^-1 in the surface soils of several dairy farms which have been in operation for up to 32 years. Soils not receiving manure had soil test P levels of only 3 mg P kg^-1. In Florida, double acid P levels of 66 mg P kg^-1 are considered high and no P additions are recommended for soils above these levels. Further, Graetz and Nair (1994) clearly demonstrated the environmental impact of P contained in these manured soils by calculating that about 4000 kg P ha^-1 would be available for transport from these areas. Also, soil test P levels (Bray-I) of up to 200 mg P kg^-1 in soils receiving long-term applications of dairy manure were observed in Wisconsin (Motshall and Daniel, 1984) and up to 279 mg P kg^-1 in soils receiving poultry litter in Oklahoma (Sharpley et al., 1993).

In many cases, manure applications have been N based, considering only soil N content and crop N requirements. This strategy can lead to an increase in soil P, due to the generally lower ratio of N:P added in manure than removed by crops (table 3). Basing manure application on soil P and crop removal of P may mitigate the excessive buildup of soil P and at the same time lower the risk for nitrate leaching to ground water. However, basing manure applications on P rather than N management could present several problems to many landowners. A strategy based on soil test P could eliminate further manure additions on much of the land area with a history of continual manure application, as several years are required for significant depletion of high soil P levels. For example, McCollum (1991) estimates that without further P addition, 16 to 18 years of cropping corn (Zea mays L.) or soybean (Glycine max (L.) Merr.) would be needed to deplete the soil test P content (Mehlich III) of a Portsmouth soil from 100 mg P kg ^-1 to the threshold agronomic level of 20 mg P kg^-1. This would force landowners to identify larger areas of land to utilize the generated manure, further exacerbating the problem of local land area limitations.

Table 3. Average P, N, and K contents (dry weight basis) of animal manures

Animal	Nitrogen	Phosphorus	Potassium
	g kg-1
Beef	32.5	5.6	2.6
Dairy	39.6	11.7	2.5
Poultry layers	49.0	20.8	2.1
Poultry broilers	40.0	16.9	1.9
Sheep	44.4	10.3	3.1
Swine	76.2	17.6	2.6
Turkeys	59.6	16.5	1.9
SOURCE: Data adapted from Gilbertson et al., 1979

Clearly, high soil test P levels are a regional problem, with the majority of soils in several states testing medium or low (figure 8). For example, most Great Plains soils still require fertilizer P for optimum crop yields. High P soils in the western States of Idaho, Utah, and Arizona may be due to high external P inputs to potato and irrigated vegetable production (figure 8). Washington and Oregon also have high indigenous geologic P levels, which can elevate soil and water P. This is not an exhaustive soil P survey, representing soils sampled by state laboratories, which thus probably involves farmers having operational nutrient management plans. However, figure 8 clearly illustrates that problems associated with high soil test P soils are aggravated by the fact that many of these soils are located near sensitive water bodies such as Florida lakes, Great Lakes, New England lakes, and Chesapeake Bay.

Amounts Transported

The main nonpoint sources contributing to the P load of water bodies are summarized in table 4. Amounts of P transported in runoff from uncultivated or pristine land are considered the background loading, which cannot be reduced. This source determines the natural status of a lake and, as will be seen later, may be sufficient to cause eutrophication. As we try to assess the impact of agricultural management on P loss in runoff, it becomes evident that little quantitative information is available on background losses of P from a given location prior to cultivation. Consequently, it is still difficult to quantify any increase in P loss following cultivation. These problems result mainly from the expensive and labor-intensive nature of water quality monitoring studies, which are site-specific and impossible to replicate, due to spatial and temporal variations in climatic, edaphic, and agronomic conditions. Despite these problems, an investigation of published studies enables generalizations regarding the effect of agricultural management on P transport in runoff.

Table 4. Nonpoint sources of P

Terrestrial sources Runoff from pristine land*** soil erosion animal excreta plant residues Runoff from cultivated land** soil erosion fertilizer loss animal excreta plant residues sewage sludge Runoff from urban land** soil erosion septic tanks domestic waste Atmosphere (cultural**; natural***) wet precipitation dry precipitation Aquatic sources Lake sediments** bottom sediments suspended sediments Biological** fauna and flora

**Difficult to control ***Impossible to control

Several surveys of U.S. watersheds have shown that P loss in runoff increases as the portion of the watershed under forest decreases and agriculture increases (figure 9). The loss of P from forested land tends to be similar to that found in subsurface or base flow from agricultural land (Ryden et al., 1973). In general, forested watersheds conserve P, with P input in rainfall usually exceeding outputs in streamflow (Schreiber et al., 1976; Taylor et al., 1971). As a result forested areas are often utilized as buffer or riparian zones around streams or water bodies to reduce P inputs from agricultural land (Lowrance et al., 1984; 1985). However, the potential loss of P from agricultural land is to a large extent dependent on the relative importance of surface and subsurface runoff in the watershed.

** Missing Image **

Figure 9. Phosphorus loss in runoff as a function of land use in the United States (adapted from Omernik, 1977).

I. Runoff

Increases in P loss in surface runoff have been measured after the application of P as fertilizer or manure (table 5). Phosphorus losses are influenced by the rate, time, and method of fertilizer application; form of P applied; amount and time of rainfall after application; and vegetative cover. The portion of applied P transported in runoff for the studies reported in table 5 was generally greater from conventionally tilled than from conservation-tilled watersheds. However, fertilizer P application to no-till corn reduced PP transport (McDowell and McGregor, 1984), probably due to an increased vegetative cover afforded by fertilization. Although it is difficult to distinguish between losses of fertilizer, manure, and native soil P, without the use of expensive and hazardous radiotracers, losses of applied P in runoff are generally less than 5% of that applied, unless rainfall immediately follows application. Also, a major portion (>75%) of annual watershed runoff can occur in one or two severe events (Edwards and Owens, 1991; Smith et al., 1991). These few events can contribute over 90% of annual P loads.

Table 5. Effect of fertilizer P application on the loss of P in surface runoff

Land use	P applied	Concentration		Amount		Fertilizer loss		Reference
		Soluble	Partic.	Soluble	Partic.	Soluble	Partic.
	kg ha-1 yr-1	mg L-1		kg ha-1 yr-1		Percent
Contour corn	40	0.19	0.71	0.12	0.45			Burwell et al. (1977), Minnesota
	66	0.25	1.27	0.15	0.76	0.1	1.2
Grass	0	0.01	0.06	0.01	0.20			McColl et al. (1977), New Zealand
	75	0.03	0.14	0.04	0.29	0.04	0.1
No till corn (silage)	0	0.23	0.46	1.10	2.20			McDowell and McGregor (1984), Mississippi
	30	0.39	0.49	0.80	1.00	0.3	+23.1+
No till corn (grain)	0	0.23	0.46	1.10	2.20
	30	0.57	0.51	1.80	1.60	2.3	+27.3+
Conventional corn	15	0.07	3.57	0.30	15.10
	30	0.11	9.71	0.20	17.50	+3.3+	16.0
Wheat-- summer fallow	0	0.30	1.80	0.20	1.40			Nicholaichuk and Read (1978), Western Canada
	54	3.70	7.40	1.20	2.90	1.9	2.8
Grass	0	0.18	0.24	0.50	0.67			Sharpley and Syers (1979), New Zealand
	50	0.98	0.96	2.80	2.74	4.6	4.1
+ Percent decrease in P loss from fertilized compared to check treatment.

II. Leaching

The loss of P in subsurface runoff is appreciably lower than that in runoff, because P is sorbed from infiltrating water as it moves through the soil profile (table 6). Subsurface runoff includes tile drainage and natural subsurface runoff; tile drainage is percolating water intercepted by artificial drainage systems, such as tile or mole drain, which accelerate its movement into streams. In general, P concentrations and losses in natural subsurface runoff are lower than in tile drainage (table 5) due to the longer contact time between subsoil and natural subsurface runoff, which enhances DP removal. Increased sorption of P from percolating water accounted for lower TP losses from 1.0 m deep (0.50 kg ha^-1 yr^-1) than 0.6 m deep (1.07 kg ha^-1 yr^-1) tiles draining a Brookston clay soil under alfalfa (Culley et al., 1983). For the shallower drains, TP losses were about 1% of fertilizer P applied, whereas 1 m deep tiles exported about 0.6% of that applied (60 kgP ha^-1 yr^-1).

Table 6. Effect of fertilizer P application on the loss of P in subsurface runoff

Land use	P applied	Concentration		Amount		Fertilizer loss		Reference
		Soluble	Partic.	Soluble	Partic.	Soluble	Partic.
	kg ha-1 yr-1	mg L-1		kg ha-1 yr-1		Percent
Alfalfa (tile drainage)	0	0.180	-	0.12	0			Bolton et al. (1970), Canada
	29	0.210	-	0.19	-	1.0	-
Continuous corn	40	0.007	-	0.03	-	-	-	Burwell et al. (1977), Iowa
	66	0.009	-	0.04	-	-	-
Terraced corn	67	0.028	-	0.17	-	-	-
Bromegrass	40	0.005	-	0.03	-	-	-
Continuous corn (tile drainage)	0	0.020	0.100	0.13	0.29			Culley et al. (1983), Canada
	30	0.110	0.360	0.20	0.42	0.2	0.4
Bluegrass sod (tile drainage)	0	0.02	0.15	0.06	0.09	-	-
	30	1.01	3.29	0.16	0.21	0.3	0.4
Oats (tile drainage)	0	0.02	0.09	0.10	0.19
	30	0.42	1.10	0.20	0.30	0.3	0.4
Alfalfa (tile drainage)	0	0.02	0.11	0.12	0.20
	30	0.37	1.03	0.20	0.31	0.3	0.3
Corn (tile drainage)	17	0.018	0.043	0.005	0.02	-	-	Hanway and Laflen (1974), Iowa
	42	0.000	0.000	0.000	0.00	-	-
	44	0.004	0.024	0.004	0.04	-	-
Grass	0	0.020	0.022	0.04	0.44			Sharpley and Syers (1979), New Zealand
	50	0.033	0.019	0.12	0.07	0.2	+7.4+
Grass (tile drainage)	0	0.064	0.072	0.08	0.09
	50	0.190	0.161	0.44	0.37	0.7	0.6
+ Percent decrease in P loss from fertilized compared to check treatment.

III. Secondary Flows

The fate and transport of P sources can occur as primary and secondary flows. The transport of P by runoff and erosion, detailed above, can be considered primary flows of P between ecosystems. Movement within and between ecosystems by more indirect means than by runoff and erosion can be considered secondary flows. Although there is little information on the relative magnitude of secondary flows of P, they may be of importance in developing sustainable agricultural systems, where we attempt to balance P inputs and outputs.

Internal secondary flows of P can occur in conservation tillage systems, where the crop residue is left in place to minimize soil water evaporation and erosion. Similarly, in other sustainable systems where cover crops are being increasingly included, the cover crop is killed before maturity and left in the soil to minimize water and light competition with the subsequent cash crop. If the crop residue is left on the surface or occasionally plowed into the soil, is the subsequent decomposition of residue P affected? Also, will the secondary flows of P within the soil environment influence the levels of P available for plant uptake and transport in runoff? Crop residue management can affect P cycling and availability as a function of residue amount, type, and degree of incorporation with tillage. Sharpley and Smith (1989) found that mineralization and leaching of residue P were greater when the residue of several crop types was surface applied rather than incorporated. A greater amount of residue will increase the amount of P being cycled and, particularly if left on the surface of the soil, will reduce evaporation losses and keep surface soil moist for more days during the growing season, thereby enhancing microbial activity and mineralization. Such secondary flows of P from crop residues can increase both the concentration and bioavailability of P transported in runoff (Langdale et al., 1985; Seta et al., 1993; Sharpley et al., 1992).

External secondary P flows include the transfer of P in grain or hay from the area of production to a confined animal operation: for example, from the Corn Belt to Texas feedlots. Thus, external secondary P flows can be considered an output from one system and an input to another. Clearly, secondary flows of P are complex as they can cross climatic, political, economic, and ecotype boundaries. Even so, they should be considered in the holistic management of P flows in terrestrial and aquatic environments.

Impacts of Phosphorus on the Environment

Terrestrial Environment

Phosphorus application has become an integral and essential part of crop production systems, providing adequate food and fiber for U.S. consumption and export demands. In addition, the judicious use and management of fertilizer P may reduce P enrichment of agricultural runoff via increased crop uptake and vegetative cover (McDowell and McGregor, 1984; Sharpley and Smith, 1991). Also, increased crop productivity resulting from P applications can allow erosive marginal lands to be taken out of productivity while maintaining yield goals.

In areas with intensive confined animal operations, the land application of manure to marginal lands as a cheap source of nutrients and organic matter has increased both grass and crop yields. This in turn can increase stocking rates for grazed pasture, as well as sales of hay to neighboring farms. Thus, carefully managed P inputs to low fertility systems can have several direct and indirect benefits to the terrestrial environment.

The negative impacts of P on the terrestrial environment are limited due to its general lack of toxicity to major cash crops and only occur once P is transported from terrestrial to aquatic environments, where eutrophication can be accelerated.

Aquatic Environment

Advanced eutrophication of surface water leads to problems with its use for fisheries, recreation, industry, or drinking, due to the increased growth of undesirable algae and aquatic weeds and oxygen shortages caused by their senescence and decomposition. Also, many drinking water supplies throughout the world experience periodic massive surface blooms of cyanobacteria (Kotak et al., 1993). These blooms contribute to a wide range of water-related problems including summer fish kills, unpalatability of drinking water, and formation of trihalomethane during water chlorination (Kotak et al., 1994; Palmstrom et al., 1988). Consumption of cyanobacterial blooms, or of water-soluble neuro- and hepatotoxins released when the blooms die, can kill livestock and may pose a serious health hazard to humans (Lawton and Codd, 1991; Martin and Cooke, 1994).

From the fisherman's point of view, advanced eutrophication of lakes can increase the population of rough fish compared to desirable game fish. This has a negative impact on the recreational value of lakes. However, fishery management often recommends a higher productivity to maintain an adequate phytoplankton-zooplankton-fish food chain for optimum commercial fish production. This food chain may be manipulated by stocking of water with certain fish species in addition to P load reductions, in efforts to reduce the incidence of algal blooms and improve overall water quality. For example, stocking lakes with predatory game fish at the top of the food chain (piscivore fish such as bass, pike, or trout) can reduce the number of planktivore or coarser fish (yellow perch, crappies), on which they feed. Zooplankton should then thrive, which in turn will reduce phytoplankton populations, improving water quality.

Reducing P inputs to lakes may not always achieve the desired or even expected water quality improvements, due to the continued contribution of P from other sources. The direct input of P in rainfall to lakes may be sufficient to enhance algal growth in certain situations. Elder (1975) estimated that rainfall P may account for up to 50% of P entering Lake Superior, and the enrichment of lakes in Ontario (Schindler and Nighswander, 1970) and Wisconsin (Lee, 1973) has been attributed to rainfall P. In addition, the release of P from sediment can sustain the growth of aquatic biota for several years after its deposition (Ahlgren, 1977; Jacoby et al., 1982; Larsen et al., 1981). Therefore, in some cases some form of in-lake management to reduce aquatic bioproductivity may be necessary and cost-effective.

Clearly, lake use has an impact on desirable water quality goals, which will require differing management. Watershed management often becomes more complex with multiple-use lakes and streams, which tend to dominate U.S. waters. For example, a reservoir may have been built for water supply, hydropower, and/or flood control, and although not a primary purpose, recreation is often considered a benefit, with aesthetic enhancement (including property value) as an additional fringe benefit.

Use-Efficiency to Decrease Negative Impacts

To increase the use-efficiency of P in agricultural systems and thereby reduce negative impacts, inputs and outputs of P in such systems must be balanced. This may be brought about by source and transport control strategies. Although we have generally been able to reduce the transport of P from agricultural land in runoff and erosion, it is clear from the extent of soils with P in excess of levels sufficient for optimum crop yields (figure 8) that we have not been as successful at minimizing soil P buildup. Strategies to minimize P loss in runoff will be most effective if sensitive or source areas within a watershed are identified, rather than widespread implementation of general strategies over a broad area. Thus, more attention should be paid to avoiding soil P buildup via management of P sources. However, before cost-effective control measures can be targeted, critical source areas vulnerable to P loss from a watershed must be identified.

** Missing Image **

Figure 10. Relationship between the growth of P-starved algae during a 29-day incubation and bioavailable P content of runoff sediment determined by ion exchange membranes and Fe-oxide strips (adapted from Sharpley, 1993).

Source Areas and In-field Targeting

I. Environmental Soil Testing for Phosphorus

Soil testing programs must measure P through the use of rapid chemical extraction procedures, if their recommendations are to be timely and cost effective. However, as we move from agronomic to environmental concerns with soils containing high P levels in excess of crop requirements, can the soil test methods developed to assess plant availability of P also estimate forms important to eutrophication? If not, are appropriate methods available? For an environmental assessment, the bioavailability of soil (or sediment) P to aquatic organisms is needed. For wastewater irrigation systems, estimates of the long-term capacity of a soil profile to retain P against leaching will be needed. Amounts of soil and sediment P extracted by P-sink approaches such as ion exchange membranes or Fe-oxide-impregnated paper strips have been shown to be closely related to the growth of P-starved algae in bioassays (figure 10; Sharpley, 1993). While these approaches are more theoretically sound than chemical extraction for estimating algal availability of runoff P, a lack of field calibration and quantification of P-sink levels limits their applicability for environmental management recommendations.

While it is unrealistic to expect that routine soil tests can provide the information needed for all environmental management programs, recent research has shown that soil test P is well correlated with several parameters needed to assess nonpoint source pollution (Sims 1993a; Wolf et al., 1985). Consequently, in areas with P-related water quality problems, soil test laboratories could use routine soil tests to provide preliminary rankings of the algal-available P content of soils (or sediment) and identify those on which the environmental test should be conducted.

The potential for soils to adsorb P is also important in the design of wastewater (manure and sewage sludge) irrigation systems, and in areas where leaching and lateral flow of P in drainage waters may be important (Sims, 1993a). The long-term capacity of soils to retain P is commonly estimated by adsorption isotherms that can be used to derive adsorption maxima for soil horizons. However, these isotherms require equilibration of soil with a series of P solutions of increasing P concentration, normally for 24 hours, and are not well adapted to routine soil test laboratories. Bache and Williams (1971), however, suggested that a single-point isotherm could be used to estimate the P adsorption maxima of soils with reasonable accuracy. This was recently confirmed by Mozaffari and Sims (1993) for surface and subsoil horizons of four Atlantic Coastal Plain soils.

Until methods to assess the environmental impact of soil P are developed, tested, and calibrated, current soil test methods will be used as surrogates. Several states have attempted to identify a soil test level where management of P in fertilizer or manure must change to reduce the potential for P loss in runoff (table 7; from Sharpley et al., 1994a). In many situations, this would require reduced or no manure and sludge applications and also the development of alternative end uses. This is a clear example of the need for an integrated approach to the use of soil test P information and the expanding role that soil test laboratories will play in this process.

Table 7. Soil P interpretations and management guidelines

State	Critical value by soil test	Management recommendation	Rationale*
Arkansas	150 mg kg1 Mehlich 3 P	At or above 150 mg kg1 STP: Apply no P from any source. Provide buffers next to streams. Overseed pastures with legumes to aid in P removal. Provide constant soil cover to minimize erosion.	CV: data from Ohio with sewage sludge. MR: reduce P levels and minimize movement of P from field.
Delaware	120 mg kg1 Mehlich 1 P	Above 120 mg kg1 STP: Apply no P from any source until STP is significantly reduced.	CV: greater P loss potential from high P soils. MR: protect water quality by minimizing further P accumulations.
Ohio	150 mg kg1 Bray P1	Above 150 mg kg1 STP: Institute practices to reduce erosion. Reduce or eliminate P additions.	CV: greater P loss potential from high P soils as well as role of high soil P in zinc deficiency. MR: protect water quality by minimizing further P accumulations.
Oklahoma	130 mg kg1 Mehlich 3 P	30 to 130 mg kg1 STP: Half P rate on slopes >8%. 130 to 200 mg kg1 STP: Half P rate on all soils and institute practices to reduce runoff and erosion. Above 200 mg kg1 STP: P rate not to exceed crop removal.	CV: greater P loss potential from high P soils. MR: protect water quality, minimize further soil P accumulation, and maintain economic viability.
Michigan	75 mg kg1 Bray P1	Above 75 mg kg1 STP: P application must not exceed crop removal. Above 150 mg kg1 STP: Apply no P from any source.	CV: minimize P loss by erosion or leaching in sandy soils. MR: protect water quality and encourage wider distribution of manures.
Wisconsin	75 mg kg1 Bray P1	Above 75 mg kg1 STP: Rotate to P-demanding crops. Reduce manure application rates. Above 150 mg kg1 STP: Discontinue manure applications.	CV: at that level, soils will remain nonresponsive to applied P for 2-3 years. MR: Minimize further P accumulations.
SOURCE: Sharpley et al., 1994a. *CV represents critical value rationale, and MR represents management recommendation rationale.

Site Vulnerability

Soil testing alone cannot assess the potential for soil P from an individual site or watershed to play a significant role in surface water eutrophication. Any environmental soil P test must be linked to site assessment of drainage, runoff, and erosion potential and management factors affecting the vulnerability for P transport from a site. For example, adjacent fields having a similar soil test P level but differing susceptibilities to runoff and erosion, due to contrasting topography or management, should not have similar P recommendations. Thus, a P indexing system was developed to identify soils vulnerable to P loss in runoff (Lemunyon and Gilbert, 1994).

 

The index is outlined in tables 8 and 9. Each site characteristic affecting P loss is assigned a weighting, assuming that certain characteristics have a relatively greater effect on potential P loss than others. The P loss potential is given a value (table 8), although each user must estab-lish a range of values for different geographic areas. An assessment of site vulnerability to P loss in runoff is made by selecting the rating value for each site characteristic from the P index (table 8). Each rating is multiplied by the appropriate weighting factor. Weighted values of all site characteristics are summed and site vulnerability obtained from table 9.

Table 8. The P indexing system to rate the potential P loss in runoff from site characteristics*

Site Characteristic (Weight)	Phosphorus Loss Potential (Value)
	None (0)	Low (1)	Medium (2)	High (4)	Very high (8)
Transport Factors
Soil erosion (1.5)**	Negligible	<10	10-20	20-30	>30
Runoff class (0.5)	Negligible	Very low or low	Medium	High	Very high
Phosphorus Source Factors
Soil P test (1.0)	Negligible	Low	Medium	High	Excessive
P fertilizer application rate (0.75)***	None applied	1-15	16-45	46-75	>76
P fertilizer application method (0.5)	None	Placed with planter deeper than 5 cm	Incorporated immediately before crop	Incorporated >3 months before crop or >3 months surface applied before crop	Surface applied <3 months before crop
Organic P source application rate (0.5)***	None applied	1-15	16-30	30-45	>45
Organic P source application method (1.0)	None	Injected deeper than 5 cm	Incorporated immediately before crop	Incorporated >3 months before crop or surface applied <3 months before crop	Surface applied >3 months before crop
*The P indexing system was developed by the following team of scientists: J. Lemunyon, D. Goss, G. Gilbert, J. Kimble, T. Sobecki, USDA-NRCS; A. Sharpley, USDA-ARS; T. Daniel, University of Arkansas; T. Logan, Ohio State University; G. Pierzynski, Kansas State University; T. Sims, University of Delaware; and R. Stevens, Washington State University. **Units for soil erosion are Mg ha-1. ***Units for P application are kgP ha-1.

Site Vulnerability	Total Index Rating Value
Low	<10
Medium	10-18
High ***Units for P application are kgP ha-1.	19-36
Very high	>36

The index is intended for use as a tool for field personnel to easily identify the agricultural areas or practices that have the greatest potential to accelerate eutrophication. It is intended that the index will identify management options available to land users and thus allow them flexibility in developing remedial strategies. Based on site vulnerability to P loss in runoff using the P index, Sims (1993b) proposed management options to minimize nonpoint source pollution of surface waters by soil P (table 10).

Table 10. Soil management options based on the P index (from Sims, 1993b)

PHOSPHORUS INDEX	Management Options to Minimize Nonpoint Source Pollution of Surface Waters by Soil P
<10 (LOW)	Soil testing: Have soils tested for P at least every three years to monitor buildup or decline in soil P. Soil conservation: Follow good soil conservation practices. Consider effects of changes in tillage practices or land use on potential for increased transport of P from site. Nutrient management: Consider effects of any major changes in agricultural practices on P losses before implementing them on the farm. Examples include increasing the number of animal units on a farm or changing to crops with a high demand for fertilizer P.
10-18 (MEDIUM)	Soil testing: For areas with low P index values, have soils tested for P at least every three years to monitor buildup or decline in soil P. Conduct a more comprehensive soil testing program in areas that have been identified by the P index as being most sensitive to P loss by erosion, runoff, or drainage. Soil conservation: Implement practices that control P losses via erosion, runoff, or drainage in the most sensitive fields. Examples include reduced tillage, wider field border strips, grassed waterways, and improved irrigation and drainage management. Nutrient management: Any changes in agricultural practices may affect P loss: consider carefully the sensitivity of fields before implementing any activity that will increase soil P. Avoid broadcast applications of P fertilizers and apply manures only to fields with low P index values.
19-36 (HIGH)	Soil testing: A comprehensive soil testing program should be conducted on the entire farm to determine fields that are most suitable for further additions of P. For fields that are excessive in P, estimates of the time required to deplete soil P to optimum levels should be made for use in long-range planning. Soil conservation: Implement practices to control P losses via erosion, runoff, or drainage. Examples include reduced tillage, wider field border strips, grassed waterways, and improved irrigation and drainage management. Consider using crops with high P removal capacities in fields with high P index values. Nutrient management: In most situations fertilizer P, other than a small amount used in starter fertilizers, will not be needed. Manure may be in excess on the farm and should only be applied to fields with low P index values. A long-term P management plan should be considered.
>36 (VERY HIGH)	Soil testing: For fields that are excessive in P, estimates of the time required to deplete soil P to optimum levels should be made for use in long-range planning. Consider using new soil testing methods that may provide more information on environmental impact of soil P. Soil conservation: Implement practices that control P losses via erosion, runoff, or drainage. Examples include reduced tillage, wider field border strips, grassed waterways, and improved irrigation and drainage management. Consider using crops with high P removal capacities in fields with high P index values. Nutrient management: Fertilizer and manure P will not be required for at least three years and perhaps longer. A comprehensive, long-term P management plan must be developed and implemented.

III. Targeting Remedial Measures

While most fresh waters are P limited (table 11), there are notable exceptions where P controls will have marginal to no benefit. In some lakes and in many streams, plant productivity is limited by high turbidity either from anthropogenic or natural sources. In others, such as the high-elevation lakes in the western United States, N is limiting (Eilers, 1991). To optimize control activities there is therefore often a need to prioritize management actions to those watersheds where the control of P will provide the greatest benefit. Management agencies are also often required to further target limited financial and human resources to those P-sensitive lakes having the highest public or ecosystem value. Phosphorus-sensitive lakes are generally those which are greater than 10 hectares in size, stratify during the summer, and have water flushing rates of less than 6 times per year. Several states are adopting a watershed approach to target nonpoint source control strategies (USEPA, 1993). The Wisconsin Department of Natural Resources (1986) targets priority lakes and watersheds by considering the threat to the water quality and the practicability of alleviating the threat; the practicability of achieving a significant reduction in P inputs; water use; and unique or endangered environmental resources.

Table 11. Limiting nutrients for various water bodies (adapted from Thomann and Mueller, 1987)

System	N/P ratio	Limiting nutrient
Rivers and streams
Point-source dominated without phosphorus removal	<<10	N
Point-source dominated with phosphorus removal	>>10	P
Nonpoint-source dominated	>>10	P
Estuaries
Freshwater region Nonpoint-source dominated	>>10	P
Freshwater region Point-source dominated	<<10	N
Brackish region	10	P or N
Saline region	<<10	N
Lakes
Large Nonpoint-source dominated	>>10	P
Small Point-source dominated	<<10	N

Remedial Strategies

I. Source Management

Efficient management of P sources on soils susceptible to P loss involves fertilizer placement and the use of soil test P recommendations based on eutrophic rather than agronomic considerations to determine P application rates. Wherever possible, subsurface placement of P away from the zone of removal in runoff will reduce the potential for P loss. Periodic plowing of no till soils may be needed to redistribute surface P accumulations throughout the root zone. Both practices may indirectly reduce the loss of P by increasing crop uptake of P and yield, which affords a greater vegetative protection of surface soil from erosion.

However, conflicts may exist within BMPs, between the Natural Resources Conservation Service guidelines for residue management and recommended subsurface applications of P. In compliance with residue conservation programs, landowners may be required to maintain a 30% residue ground cover. Under this BMP, subsurface application or knifing of P fertilizer or manure, which may be recommended to minimize P loss in runoff, could be unacceptable if it reduces residue cover below 30%. Thus, BMPs should be flexible enough to enable modified residue and P management plans to be compatible.

Is it then possible to select a "scavenger" crop that may have a higher affinity or requirement for P and thereby reduce soil nutrient stratification? Alfalfa, for example, has reduced subsoil nitrate accumulations (Mathers et al., 1975). May the same be true for surface soil accumulations of P? It is possible that by utilizing residual soil P, careful crop selection will reduce the amount of nutrients potentially available to be transferred to surface waters (Pierzynski and Logan, 1993).

II. Transport Management

Loss by erosion and runoff may be reduced by increasing vegetative cover through conservation tillage. However, DP and BAP losses can be greater from no till than from conventional till practices. Accumulation of crop residues and added P at the soil surface provide a source of P to runoff that would be decreased during tillage. Such water quality tradeoffs must be weighed against the potential benefits of conservation measures in assessing their effectiveness.

Additional measures to minimize P loss by erosion and runoff include buffer strips, riparian zones, terracing, contour tillage, cover crops, and impoundments or small reservoirs. However, these practices are generally more efficient at reducing PP than DP. For example, several studies have indicated little decrease in lake productivity with reduced P inputs following implementation of conservation measures (Gray and Kirkland, 1986; Young and DePinto, 1982). The lack of biological response was attributed to an increased bioavailability of P entering the lakes as well as internal recycling. Clearly, effective remedial strategies must address the management of P sources and applications as well as erosion and runoff control.

Although P losses in runoff are generally < 5% of applied P, DP and TP concentrations often exceed critical values associated with accelerated eutrophication (0.05 and 0.1 mg L^-1; USEPA, 1976; Vollenweider and Kerekes, 1980). This is true even for unfertilized native grass watersheds (Sharpley et al., 1986). Also, P inputs in rainfall can contribute to freshwater eutrophication (Lee, 1973; Schindler, 1977). Thus, the measures recommended above may not reduce P loss in runoff from cultivated land to critical values. This emphasizes the need to target remedial measures on source areas where the potential for P loss is greatest. Further, the critical level approach should not be used as the sole criterion in quantifying permissible tolerance levels of P loss in runoff as a result of differing management. A more flexible approach advocated by limnologists considers the complex relationships between P concentration and physical characteristics of affected watersheds (runoff and erosion) and water bodies (mean depth and hydraulic residence time) on a site-specific and recognized-need basis.

III. Economic Management

One of the major obstacles facing landowners is overcoming the economic restrictions of moving manure to a greater acreage, where it could supplement or even replace mineral fertilizer requirements. The recent trend in the formation of cooperatives that can more cost-effectively compost and compact manure should be encouraged by cost-sharing programs. Neighboring landowners and private industry are also developing manure processing alternatives. Examples of this include centralized storage and distribution networks, regional composting facilities, and pelletizing operations that can produce a value-added processed manure for distribution to other areas. Pelletization is a particularly attractive option because it results in a dried, lightweight material that can be handled, transported and applied in much the same manner as commercial fertilizer.

By composting and compacting, the bulk density of the manure is reduced, as is the cost of transportation. If the consumer is not willing to bear a part of the financial support, then it may be necessary to recommend producers and landowners to take part in cooperative manure treatment programs. The level of involvement could be linked to the number of animals per farm.

Storage of manure will allow more flexibility in timing applications. A wide range of storage methods and costs are available to landowners (Brodie and Carr, 1988). For some solid manures, inexpensive plastic sheeting can perform well with very low cost. However, all storage methods must be managed carefully to fully realize their potential in an agronomically and environmentally sound BMP.

Another economic option is the buying and selling of pollution credits within a given watershed, similar to that recently adopted for air quality control. Farmers able to limit P loss below recommended levels could sell credits to a farmer unable to meet these levels. The number of credits a farmer has could be linked to the number of animals and area of farm. As a result, P export from a watershed may be kept within predetermined limits by sharing the responsibility among farmers. Education and extension programs should highlight the nutritive and mulching value of manure to non-producing farmers: in effect, increasing the demand for this nutrient resource.

Even so, it is clear that our current technology will not permit an unlimited number of animals in a region without impacts on water quality. Thus, it may be necessary to redistribute animals or to limit animal numbers within an area. Several states now require that new animal facilities which exceed a certain size have an appropriate waste management plan. Thus, it is essential that we develop and transfer technology to implement environmentally sound recommendations for manure management.

Conclusions

The fate of P in soil is governed by dynamic climatic, edaphic, and agronomic variables. Up to 80% of P applied to soil can react with Al, Fe, and Ca to form complexes that are unavailable for plant uptake. This P can, however, be transported from the site of application by runoff and erosion. Unless added P is incorporated into the soil, it usually accumulates in the surface 10 cm of soil, increasing the potential for its transport in runoff.

Although we have been successful in reducing P inputs to aquatic systems via point sources, municipal and urban discharge, and detergents, less success has been achieved in minimizing nonpoint agricultural inputs. This is exacerbated where P input in manure from confined animal operations often exceeds local crop removal rates. The subsequent accumulation of P in soil is of environmental rather than agronomic concern in many cases. As many years are required to bring about a significant reduction in soil P levels by crop removal, time is not on our side. Also, once lake eutrophication is accelerated, it is usually not cost effective to treat the lake, and internal recycling of sedimentary P can support the growth of aquatic biota even if external inputs could be stopped.

Consequently, efforts to minimize P transport from terrestrial to aquatic environments and to slow down freshwater eutrophication must identify critical source areas of P in a watershed that present a greater risk to P-sensitive lakes, in order to target cost-effective remedial strategies. In areas of confined animal operations, the development and adoption of innovative measures to transport manure greater distances and to find alternative end-uses must be encouraged. Finally, perhaps most crucial to any strategy for water quality improvement is efficient transfer of research technology to the land user. Effective implementation will involve education programs to overcome the perception by end-users of water that it is often much cheaper to treat the symptoms of eutrophication rather than control the nonpoint sources.

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