ADVANCES IN AGRONOMY, VOL. 41

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS

ON THE ENVIRONMENT A. N. Sharpley and R. G. Menzel

Water Quality and Watershed Research Laboratory United States Department of Agriculture (USDA)

Agricultural Research Service (ARS), Durant, Oklahoma 74702

I. INTRODUCTION

Phosphorus in the form of phosphate (P) is essential for plant growth, and its application to agricultural land often improves crop production. Production per unit area is increased with fertilizer P, The increased plant cover that is possible with proper use of fertilizer can reduce soil erosion from the cultivated area. Addition of phosphorus to fish ponds may also in- crease fish production. Although P is not toxic, the continued application of P fertilizer can produce detrimental effects in both the terrestrial and aquatic environments. In any discussion of the impact of soil and fertilizer P on the environment, the detrimental effects must be considered along with the benefits .

Although there are no direct detrimental effects of P on the terrestrial en- vironment, the continued application of fertilizer P to agricultural land can result in the buildup of natural trace contaminants contained in the fertilizer. There have been concerns, for example, about the accumulation of the con- taminants cadmium, uranium, and radium in P-fertilized soils. The transport of P from the terrestrial to aquatic environment in surface and subsurface runoff can result in a deterioration in water quality from accelerated eutrophication. Eutrophication of surface water leads to problems with its use for fisheries, recreation, industry, or drinking, due to the increase in growth of undesirable algae and aquatic weeds. In addition to P, nitrogen (N) and car- bon (C) are also commonly associated with these problems. However, control of accelerated eutrophication through limiting C and N inputs is restricted, due to the difficulty in controlling atmospheric exchanges of these elements. Thus, P is often the limiting element and its control is of prime importance in reducing the accelerated eutrophication of surface waters. It should be recognized, however, that even where the most severe eutrophication

297

298 A. N. SHARPLEY AND R. G . MENZEL

problems arise, there is no fundamental threat to the general environment, although fish kills can occur due to oxygen depletion in a water body. As Taylor and Kilmer (1980) pointed out, the problems are due to biological productivity of the aquatic environment when a lower level is desired for aesthetic, recreational, or economic reasons.

Only part of the P in the terrestrial and aquatic environment is available for plant growth. While nearly all of the soluble inorganic P (Pi) is readily available, much of that attached to eroded soil or sediment (particulate P) may be unavailable. Soluble organic P (Po) may become available through hydrolysis and phosphatase enzyme mineralization. As exchange between the soluble and particulate forms of P can occur in both the terrestrial and aquatic environments, knowledge of these processes is essential in eval- uating the impact of soil and fertilizer P on the environment.

This chapter discusses the impacts of soil and fertilizer P on the terrestrial environment, the processes involved in the transport of P from the ter- restrial to the aquatic environment, and finally, the impact of P on the aquatic environment.

II. IMPACT OF PHOSPHORUS ON THE TERRESTRIAL ENVIRONMENT

A. BENEFICIAL EFFECTS

Fertilizer P use has become an integral and essential part of the food production system. Its use permits adequate food and fiber production for domestic consumption and export demands (Viets, 1975). In addition, fertilizer use permits this production on a reduced acreage, thereby benefiting the environment in several ways. If production is confined to a smaller acreage, less herbicides and insecticides are needed. Finally, if the more erodible land can be kept in grass or forest cover, erosion is minimized. Thus, a larger proportion of the terrestrial environment will have less disturbance, and less sedimentation will occur in aquatic environments.

Fertilizer P has been used to establish a vegetative cover on infertile and badly eroded soils, which reduces the transport of water, soil, and nutrients in surface runoff. The fact that fertilizers can have a positive effect as a conser- vation tool to minimize runoff was recognized in New Zealand 30 years ago (Campbell, 1950). The role of P fertilizers in forest establishment, regenera- tion, and production on several soils has been demonstrated (Baule, 1973). Thus, from an economical and environmental standpoint, P fertilizers have an important beneficial function.

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 299

B. DETRIMENTAL EFFECTS

Potentially toxic chemical elements may be introduced into the food chain by adding P fertilizer to the soil (Tremearne and Jacobs, 1941; Bowen, 1966; Lisk, 1972). This results from the fact that several heavy metals, such as arsenic, cadmium, chromium, lead, and vanadium, occur in P rock ore and are not eliminated during the manufacture of P fertilizer.

The cadmium (Cd) content of P fertilizers has been studied extensively, due to its common occurrence in P rock, long-term persistence in the soil, uptake and accumulation by plants and animals, and toxicity at low levels (Schroeder and Balassa, 1961; Lagerwerff, 1971; Mortvedt, 1987). The Cd content of P fertilizer has been shown to vary with the source of the P rock and concentration of P in the fertilizer (Mortvedt and Giordano, 1977). Phophorus fertilizers produced from Florida deposits generally have a lower Cd content (10-20 mg Cd/kg) than those from western U.S. deposits (50-200 mg Cdlkg). Williams and David (1973) found that Australian P fertilizers contained from 25 to 50 mg Cd/kg. In addition, they observed that P and Cd contents of the fertilizer material were highly correlated and suggested that most of the Cd in P rock was concentrated in phosphoric acid during the manufacture of high-analysis fertilizers.

Several studies have reported an increase in Cd content of soil in the cultivated layer following application of high rates of P fertilizer (Table I). At normally recommended fertilizer P rates, however, little Cd accumulation has been found in crops following long-term applications (> 50 years) (Mortvedt, 1987). The accumulations of Cd in soil have resulted in increases in the Cd con- tent of certain plants (Table I). Schroeder et al. (1967), however, found that the Cd concentration in many plant species did not increase and in several species decreased as a result of fertilizer P application. Apparently, differences in Cd uptake occur between plant species. Furthermore, the bioavailability of Cd in- creases with a reduction in soil pH (CAST, 1976; Williams and David, 1976; Mortvedt et al., 1981) due to a decrease in Cd sorption on the soil (Anderson and Nielson, 1974; Garcia-Miragaya and Page, 1978; Jarvis and Jones, 1980). Consequently, long-term production of Cd-accumulating crops on acid soils (PH 5.5) may require special P fertilizers with low Cd content.

No detectable increase in arsenic, chromium, lead, or vanadium concen- tration in soil was found following the application of 8888 kg/ha of concen- trated superphosphate (Goodroad and Caldwell, 1979). It is unlikely that there is any danger of contamination following P fertilization as long as the content of these elements in P fertilizers remains low. Similarly, Mortvedt and Giordano (1977) concluded that the plant uptake of chromium and lead in fertilizer was not significant at the rates of P usually used.

In addition to heavy metals, P fertilizers contain radioactive material from the rock source in amounts between 30 and 200 mg/kg uranium (v)

Table I

Effect of P FertWzer Application on the Amonnt of Cadmiam in Soils and Plants

Cadmium content of

cadmium soid Plant content of Fertilizer

P applied Duration fertilizer Unfertilized Fertilized Unfertilized Fertilized Reference Description WhdYr) (yr) ( m g W (mgflrg) (mglkg) (mg/kg) ( m g k s )

Andersson and hlahlin (1981) Clay: barley grain 5 15 23 0.23 0.022' 0.013 0.013' 15 15 23 0.23 0.023' 0.013 0.013' 45 15 23 0.23 0.024' 0.013 0.016

Fine sand: barley 5 15 23 0.07 0.06' - - 15 15 23 0.07 0.070 - - 45 15 23 0.07 0.W - -

Andrews ef of. (1979) Clay loam: grass/clover 40 30 5 0.02 0.04 0.003 0.005

Mortvedt er of. (1981) Sit loam: wheat grain 50 2 153 0.07 0.11 0.03 0.09 straw 50 2 153 0.07 0.11 0.07 0.12

Mulla ef of. (1980) Sandy loam: barley 175 36 174 0.7 1 .o 0.01 0.01' Swiss chard 175 36 174 0.7 1 .o 0.26 1.60

Reuss ef of. (1978) Silt loam: raddish 127 1 174 - - 0.40 3.40 lettuce 127 1 174 - - 0.20 6.30 peas 127 1 174 - - 0.20 0.90

Williams and David (1973, Krasnozem: oats 125 20 50 0.05 0.45 0.03 0.28 1976) clover 125 20 50 0.05 0.45 0.10 1.07

lucerne 125 20 50 0.05 0.45 0.09 0.46

'No significant effect at 5% level.

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 301

and 10 mg/kg thorium (Caro, 1964; Barrows, 1966; Menzel, 1968). Thus, the addition of radionuclides with high application rates of superphosphate for a century is similar to amounts occurring naturally in the plow layer. Marsden (1964) reported that topdressing of pasture with superphosphate for 16 years at the high annual rate of 2200 kg/ha resulted in only a 5 % in- crease in a activity in the soil. In a recent investigation of U accumulation from the long-term application of superphosphate (33 kgP/ha/yr) to a clay loam under pasture at Rothamsted, England, Rothbaum et al. (1979) observed that most of the U applied since 1889 (1.3 kg/ha) was retained, like P, in the plow layer. The radiation hazard which might result from the uptake of radionuclides into food plants from applied fertilizer appears to be negligible (Menzel, 1968; Mays and Mortvedt, 1986). The production of phosphoric acid removes almost all radioactive contaminants. Thus, high- analysis P fertilizers have no radiation hazard.

Due to similar strengths of P and U sorption by soil, Menzel(l968) sug- gested that losses of U from surface soil could occur by erosion and might be similar to the losses of added fertilizer P. In addition, U is generally con- sidered to be mobile in the absence of organic matter (Hostetler and Gar- rels, 1962; Schultz, 1965) and may, thus, leach from sandy soils containing little organic matter. In fact, a significant increase in the U content of rivers draining intensively fertilized and farmed agricultural land in the southwestern United States was measured by Spalding and Sackett (1972). The U increase was in some cases attributed to the application of P fer- tilizer.

The heavy metal and radionuclide contaminants discussed are generally strongly absorbed by soil, as is P. Consequently, these contaminants may be preferentially transported with finer soil particles during rainfall and ero- sion and accumulate in deposited sediment.

111. TRANSPORT OF PHOSPHORUS FROM THE TERRESTRIAL TO AQUATIC ENVIRONMENTS

The transport of P from terrestrial to aquatic environments in runoff can occur as either soluble or particulate P. The term particulate P includes P sorbed by soil particles and organic matter eroded during runoff. Soil ero- sion is a selective process in which runoff sediment becomes enriched in finer-sized particles and lighter organic matter. Because P is strongly ab- sorbed on clay particles (Syers et al., 1973a; Barrow, 1978; Parfitt, 1978; Sibbesen, 1981) and organic matter contains relatively high levels of P, the major proportion of P transported to the aquatic environment from cultivated land is usually in the particulate form (Burwell et al., 1977;

302 A. N. SHARPLEY AND R. G. MENZEL

Logan et al., 1979; Nelson et al., 1979; Sharpley and Syers, 1979). In runoff from grassland or forest soils, which carries little suspended soil, most of the P may be transported in the soluble form (Burwell et al., 1975; Singer and Rust, 1975).

Most soluble P forms found in runoff are biologically available, but the bioavailability of particulate P from various sources differs greatly (Syers et al., 1973b; Porter, 1975; Lee et al., 1978; McCallister and Logan, 1978; Logan et al., 1979). In addition, transformations between the two P forms can occur during transport (Carter et al., 1971; Kunishi et al., 1972; Sharpley et al., 1981~). Consequently, knowledge of the mechanisms in- volved in the extraction and detachment of soluble and particulate P during runoff, in addition to knowledge of the nature of the particulate matter in runoff and the various sources and amounts of P, is important in evaluating the impact of soil and fertilizer P on the aquatic environment.

A. AMOUNTS TRANSPORTED FROM TERRESTRIAL ENVIRONMENTS

Increases in the amounts of soluble and particulate P transported in sur- face runoff have been measured after the application of fertilizer P (Table 11). These increases result from an increase in the available P content of sur- face soil (Barrow and Shaw, 1975; Elrashidi and Larsen, 1978; Fukely, 1978; Barber, 1979) and total P content of eroded soil material, respective- ly, compared to unfertilized soil. The losses of fertilizer P are influenced by the rate, time, and method of fertilizer application; form of fertilizer; amount and time of rainfall after application; and vegetative cover. Detail- ed reviews of the effect of fertilizer P on the amounts of P transported from agricultural land have been presented previously (Ryden et al., 1973; Viets, 1975; Timmons and Holt, 1980). Though it is difficult to distinguish be- tween losses of fertilizer P and native soil P, the losses of fertilizer P are generally less than 1% of that applied. The losses of P in subsurface drainage are small, with applications of fertilizer at recommended rates nor- mally having no significant effect on P losses.

Phosphorus losses in surface runoff may be reduced by incorporating fer- tilizer material into the surface soil away from the zone of extraction and detachment and by using conservation or minimum tillage methods to reduce soil erosion. The two main consequences of conservation tillage are the increase in amount of residues on the surface and the reduction in mechanical manipulation and mixing of the soil. Although this may result in decreased runoff volumes (Burwell and Kramer et al., 1983; Langdale et al., 1983; McDowell and McGregor, 1984; Moldenhauer et al., 1983; Wendt and Burwell, 1985), P can build up in the surface 0-3 cm of soil (McDowell and McGregor, 1984; Randall, 1980; Wells, 1985). Consequently, the in- teraction between runoff water and surface soil and subsequent transport of

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 303

soluble and particulate P in runoff is affected. The loss of total P in runoff from conservation tillage practices is lower than from conventional prac- tices, due to reduced erosion rates and runoff volumes with the former (An- draski et al., 1985; Langdale et al., 1985; McDowell and McGregor, 1984). In contrast, slight increases in soluble P loss from no-till land with unincor- porated residues compared to incorporated residues have been reported (Romkens et al., 1973; Barsisas et al., 1978; Reddy et al., 1978; Langdale et al., 1985). For example, McDowell and McGregor (1984) found that conserva- tion tillage (no-till) reduced total P losses nine-fold compared to conventional practices for corn (for silage) in Mississippi. However, an eight-fold increase in soluble P loss was measured for the no-till compared to conventional tillage.

Conservation tillage practices have been shown to increase the proportion of clay-sized particles transported in runoff and, thus, increase the P:sedi- ment ratio (Logan and Adams, 1981). In a recent study, Andraski et al. (1985) reported a 63% reduction in algal available P (resin extractable) loss in runoff from no-till compared to conventionally tilled corn. This highlights the need to consider the algal availability (or bioavailability) of particulate P transported in runoff when evaluating management decisions aimed at reducing the impact of P on the trophic status of a waterbody. In an attempt to improve the trophic status of Lake Erie, The U.S. Army Corps of Engineers (1982) showed that nonpoint-source particulate P load needed to be reduced by about 26% to reduce eutrophication significantly. From field experimentation and calculation, Forster et al. (1985) concluded that accelerated implementation of conservation tillage, including no-till, on soils suited to those practices on the United States side of Lake Erie basin can achieve the required particulate P load reduction in 20 years.

Even though the amounts of P transported in runoff may be small com- pared with amounts applied, it is evident that P concentrations of both sur- face and subsurface runoff are greater than the critical values (0.01 and 0.02 mg/liter for soluble and total P, respectively) suggested by Sawyer (1947) and Vollenweider (1968), above which biological growth can be stimulated. This is also true for several unfertilized watersheds (Table 11). In addition, P levels in rainfall may exceed the critical values (Schindler and Nighswander, 1970; Murphy and Doskey, 1975; Sharpley et al., 1985b; Tabatabai et al., 1981) and can result in natural eutrophication (Schindler, 1977; Lee, 1973). Consequently, the critical P level approach should not be used as the sole criterion in quantifying permissible tolerance levels of P in surface runoff as a result of differing management practices (Sharpley et al., 1985a).

B. PHOSPHORUS DESORPTION

The first step in the transport of soluble P is the desorption and dissolu- tion of P from soil. The desorption of P from soil material in relation to

Table Il

Effect of P Fertilization on the Concentration and Amounts of P Trnnsported in Surface Runoff and Subsurfnce Drainage

Concentration Amount P

applied Soluble P Particulate P Soluble P Particulate P Reference and location Land use (kg/ha/yr) (mghter) (mg/liter) (kg/ha/yr) (kg/ha/yr)

Surface runoff

Burwell et u/. (1975). Minnesota Fallow Hay Contour corn Rotation corn Rotation oats

Burwell el a/. (1977), Iowa

McColl et ul.

McDowell et a/.

(1977), New Zealand

(1980), Mississippi

Contour corn Contour corn Grazed bromegrass Terraced corn

Native forest Pasture

Corn grain Corn silage

0 0

29 29 30

66 40 41 67

0 75

30 30

0.25 0.19 0.80 0.57

0.01 0.03

0.11 0.05

- 1.27 0.71 0.40 1.14

0.06 0.14

4.5 4.4

0.10 0.39 0.25 0.15 0.14

0.15 0.12 0.16 0.10

0.01 0.04

4.3 0.2

33.15 0.02

18.19 8.43 5.01

0.76 0.45 0.08 0.20

0.20 0.29

0.02 0.02

w 0 v,

Menzel et 01. (1978). Oklahoma Rotation grazing Wheat Cotton

Nicholaichuk and Read (1978), Wheat/ western Canada summer fallow

Romkens and Nelson (1974), Indiana

Fallow

Sharpley and Syers Pasture (1979), New Zealand

Subsurface drainage

Bolton el 01. (1970), Canada

Burwell el al. (1977), Iowa

Hanway and Laflen (1974), Iowa

0 0.02 0.6 0.04 0.50 6.5 0.28 0.75 0.30 1.90

25 0.72 1 .00 1.10 5.60

0 0.3 1.8 0.20 1.40 54 3.7 7.4 1.20 2.90 0 0.07 -

56 0.24 - - - 113 0.44 - - -

- -

0 0.20 0.24 0.50 0.67 50 0.98 9.61 2.80 2.74

Alfalfa 0 0.180 - 0.12 - (tile drainage) 29 0.210 - 0.19 -

Continuous corn 66 0.009 - 0.04 - Continuous corn 40 0.007 - 0.03 - Grazed bromegrass 41 0.005 - 0.03 - Terraced corn 67 0.028 - 0.17 - Corn 38 0.018 - 0.005 -

(tile drainage) 96 O.Oo0 - O.Oo0 - 100 0.004 - 0.004 -

Sharpley and Syers Pasture 0 0.020 - 0.004 - (1979), New Zealand 50 0.033 - 0.12 -

Pasture 0 0.064 - 0.08 - (tile drainage) 50 0.190 - 0.44 -

306 A. N. SHARPLEY AND R. G. MENZEL

plant availability and water quality has been studied using various extrac- tion mediums and so1ution:soil ratios (Table 111). Few studies have used filtered runoff (Wang, 1974) or lake water (Bahnick, 1977) as the support medium, due to the technical problems involved in preparing large volumes of filtrate of constant chemical composition. Bahnick (1977) reported an in- crease in P desorption from clay deposits from Lake Superior to deionized water compared to filtered Lake Superior water, which was attributed to a lower pH of the deionized water. As the amount of P desorbed depends upon the ionic strength and cationic species in extracting medium (Ryden and Syers, 1977b) and so1ution:soil ratio used (Hope and Syers, 1976; Bar- row and Shaw, 1979; Sharpley et al., 1981b), a need for the standardization of methodologies used to relate P desorbed to the potential availability of P to plants and transport in runoff is indicated. These conditions should be related to the soil solution composition, simulating field conditions as close- ly as possible (Wendt and Alberts, 1984). Experimental conditions should

Table I11

Methods Used to Determine Desorbable P

Solution: soil Extractant ratio References

0.1 M NaCl

0.01 M CaCI,

Anion exchange resin (Dowex I-X4)

(Dowex I-XB) (Dowex bX4) (Dowex 21K)

Distilled water

j*P + distilled water j2P + 0.1 M NaCl I*P + 0.1 M NaCl jrP + stream water

Filtered lake water (Lake Superior)

100: 1 50 : 1 50 : 1

10: 1 1 0 : 1 1 0 : 1 5 : 1

25 : 1 6 : 1-300 : 1

5 0 : 1

30 : 1 100: 1 25 : 1 50 : 1

10 : 1-1000 : 1 2000 : 1-4000 : 1

1 : l 1 0 0 : 1 40: 1

100: 1 - 1 m : I

Li et at. (1972) Ryden et at. (1972) Romkens and Nelson (1974)

White and Beckett (1964) Taylor and Kunishi (1971) Gardner and Jones (1973) Elrashidi and Larsen (1978) Green el a/. (1978) Barrow (1979) Oloya and Logan (1980)

Ballaux and Peaslee (1975) Evans and Jurinak (1976) Vig ef at. (1979) Bache and Ireland (1980)

Sharpley et al. (1981b) Bahnick (1977)

Baker (1964) Li el a/. (1972) Ryden and Syers (1977a) Schreiber et a/. (1977)

2000: 1-4OOo: 1 Bahnick (1977)

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 307

be governed by soil type and processes simulated, be it release to soil solu- tion (narrow so1ution:soil ratio) or runoff (wide so1ution:soil ratio).

The desorption of soil P is rapid (Kunishi et al., 1972; Ryden and Syers, 1977b; Oloya and Logan, 1980). Evans and Jurinak (1976) reported that 50% of P desorbed from a desert soil in 50 hr occurred in the first hour of reaction, and Sharpley et al. (1981b) found that approximately 75% of the P desorbed in 4 hr from several soils occurred in the initial 30 min. Conse- quently, P can be desorbed from surface soil by short rainfall and runoff events. In fact, a highly significant linear relationship between the amount of desorbable P in the surface soil and the soluble P concentration of sur- face runoff has been found (Hanway and Laflen, 1974; Romkens and Nelson, 1974; Sharpley et al., 1978, and 1981a). In the case of soils contain- ing low amounts of readily desorbed P (10-90 mg P/kg), Oloya and Logan (1980) observed that the pool of desorbable P was large compared to the amount that readily desorbed (5-8Vo of the desorbable P in 24 hr). Conse- quently, these soils may release low levels of P over a long period of time.

The desorption of soil P by rainfall runoff water is brought about by in- teraction with a thin layer of surface soil (1-3 mm) (Donigian et al., 1977; Ahuja et al., 1981; Sharp’ ~ et al., 1981a; Sharpley, 1985a). If the surface water percolates through the il profile, sorption of P by P-deficient sub- soils generally results in low concentrations of soluble P in subsurface flow (Ryden et a/., 1973; Baker et al., 1975; Burwell et al., 1977; Sawhney, 1977; Sharpley and Syers, 1979). 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 A1 (Fox and Kamprath, 1971; Hortensteine and Forbes, 1972; Singh and Jones, 1976; Duxbury and Pever- ly, 1978; Miller, 1979). Similarly, P is more susceptible to movement through sandy soils with low P sorption capacities (Ozanne et al., 1961; Adriano et al., 1975; Sawhney, 1977) and in soils which have become waterlogged, where a decrease in Fe (111) content occurs (Ponnamperuma, 1972; Gotoh and Patrick, 1974; Khalid et al., 1977).

c. PHOSPHORUS LEACHED FROM VEGETATION

The transport of P from the terrestrial to aquatic environments may oc- cur through the leaching and washoff of P from growing and decaying plant material (Gburek and Broyan, 1974; McDowell et al., 1980; Schreiber and McDowell, 1985; Schreiber, 1985; Sharpley, 1981). Numerous studies have suggested that the leaching of vegetation in different stages of growth and decay may account in part for the seasonal fluctuations in soluble P transported in runoff from various watersheds (Kleusner, 1972; Wells et al., 1972; Gosz et al., 1973; White and Williamson, 1973; Burwell et al., 1975; McDowell et al., 1980). Similarly, Muir et al. (1973), Burwell et al. (1974),

308 A. N. SHARPLEY AND R. G. MENZEL

and Gburek and Heald (1974) attributed differences in amounts of P transported to differences in the type of vegetation from watershed to watershed. Increased losses of soluble P in runoff from alfalfa plots (33 g P/ha) compared to forested (4 g Piha), oats (16 g P/ha), and corn plots (1 1 g P/ha), were attributed to the larger amounts of P leached from alfalfa (Wendt and Corey, 1980). In addition, White et al. (1977) reported that nutrient transport in surface runoff was significantly related to runoff amounts and that 50-75070 of the data variation was explained by this fac- tor. The unexplained variation was partially attributed to plant cover, growth, and stage of decomposition.

The extraction of cut vegetation by deionized water has shown that large differences in the amounts of P leached can occur with vegetation type and that they increased dramatically following freezing and thawing of the vegetation (Timmons et al., 1970; Bromfield and Jones, 1972; Cowen and Lee, 1973; White, 1973). Using a multiple-intensity rainfall simulator, Schreiber and McDowell (1985) measured 125 g P/ha (6.3% of wheat P) leached from wheat straw residue (4500 kg/ha loading rate) during a 25-mm rainfall of 25 mm/hr. Although the P losses generally increased as wheat residue loading rate increased, the percentage of P removed from the wheat residue decreased, due possibly to an easier pathway of nutrient movement from the residue surface to runoff at lower residue loadings (Schreiber, 1985). In one study, Sharpley (1981) found that growing cotton, sorghum, and soybean plants could maintain concentrations of soluble P in plant leachate (0.018-0.154 mg P/liter) similar to those released from unfertilized soil. For healthy mature plants, leached P accounted for approximately 20% of soluble P transported in surface runoff. Under conditions in which plants became P deficient and senescing plants, however, canopy leachate contributed the major proportion (90%) of the soluble P transported in sur- face runoff. Depending on the relative rates of infiltration and runoff, a portion of plant leachate may infiltrate the soil and be recycled.

D. DETACHMENT AND TRANSPORT OF PARTICULATE PHOSPHORUS

Sources of sediment and particulate P in streams include eroding surface soil, subsoil, and stream sediments derived from streambanks and channel beds. The primary source of sediment in watersheds with a permanent vegetative cover, such as forest or pasture, is from streambank erosion. This sediment will have characteristics similar to the subsoils or parent material of the area, which are often P deficient. In cultivated watersheds, however, streambank erosion constitutes a smaller proportion of the sedi- ment eroded, due to surface soil erosion. Although attempts to identify sediment sources have been made using sediment mineralogy (Klages and Hsieh, 1975; Wall and Wilding, 1976; Sawhney and Frink, 1978), erroneous

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 309

conclusions can be made if preferential erosion of clay minerals occurs during transport (Jones et al., 1977; Murad and Fischer, 1978; Rhoton et al., 1979).

The degree of enrichment of P in runoff sediment due to the preferential transport of finer-sized particles and lighter organic matter discussed earlier is expressed as an enrichment ratio. For P, this is calculated as the ratio of the concentration of P in the runoff sediment to that in the source soil. Enrichment ratio values of 1.3 for total P and 3.3 for 0.001 M H,S04 ex- tractable P for a silt loam situated on a 20-25% slope were observed by Rogers (1941), whereas Massey and Jackson (1952) observed values between 1.9 and 2.2 for “water soluble plus pH 3 extractable” P for silt loams in Wisconsin. Other ER values have ranged from 1.5 to 8.9 for total P (Knoblauch et al., 1942; Neal, 1944; Stoltenberg and White, 1953; Duffy et al., 1978). Sharpley (1985b) observed that the enrichments of Bray-1 (2.45) and labile P (2.89) were greater than those for other forms of P (total, in- organic and organic) (1.48) for six soils using simulated rainfall. The relatively greater enrichment of available P forms was attributed to less ag- gregation of runoff sediment compared to source soil reducing the physical protection of P. Phosphorus desorption-sorption characteristics, buffer capacity, sorption index, and equilibrium P concentration (EPC) were also enriched in runoff sediment compared to source soil.

Massey and Jackson (1952), Menzel(1980), and Sharpley (1980) reported that the logarithm of particulate P ER was linearly related to the logarithm of soil loss (kg/ha). A similar relationship between soil loss and the ER of labile (”P) and bioavailable P (0.1 M NaOH) and of P sorption-desorption characteristics (P buffer capacity, P sorption index, and EPC) was also measured for several soils under simulated rainfall (Sharpley, 1985b). Although the texture of the soils studied ranged from Bernow fine sandy loam (8% clay) to Houston Black clay (50% clay), regression equations of the logarithmic soil loss-enrichment ratio relationship were similar. Dif- ferent regression coefficients were obtained for different nutrient forms, however, with equation (1) holding for bioavailable, Bray 1, total, par- ticulate, and organic P, P buffer capacity, and P sorption index:

In ER = 1.21 - 0.15 In Soil loss (kg/ha) for labile P: In ER = 2.48 - 0.35 In Soil loss and for EPC: In ER = 1.63 - 0.25 In Soil loss

(1) (2) (3)

The development of these general relationships between ER and soil loss provide a method of estimating the transport of these P forms with sedi- ment (Sharpley, 1985b). Inclusion of these relationships in water quality models will improve the estimation of biological productivity of surface water in response to inputs of nutrients from agricultural runoff and allow a better description of P-runoff sediment interactions.

3 10 A. N. SHARPLEY AND R. G. MENZEL

E. CHANGES BETWEEN PHOSPHORUS FORMS DURING TRANSPORT

Interchange between soluble and particulate P can occur during transport in stream flow. 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 short-term potential of runoff to increase algal growth. In addition, soluble P may be removed by stream macrophytes (Stake, 1968; McColl, 1974; Vincent and Downes, 1980) and particulate P deposited or eroded from the stream bed with a change in stream velocity. Thus, the amounts of soluble and particulate P entering lakes and impoundments can be quite different from those entering stream flow.

The direction of the exchange between soluble and particulate forms will depend on the concentrations of sediment and soluble P in stream flow and the equilibrium P concentration of the sediments contacted, which will in- clude suspended, stream bank, and bottom material. The extent of these changes will depend on the labile or desorbable P content of the sediment material contacted and rate of stream flow.

The equilibrium P concentration (EPC,) is defined as the soluble P con- centration that is supported by a solid sample at which no net sorption or desorption takes place (White and Beckett, 1964; Taylor and Kunishi, 1971). If the soluble P concentration of runoff or stream flow falls below the EPC, of the suspended stream bank material contacted, P will be desorbed from the material. If, however, the soluble P concentration in- creases above the EPC,, P will be sorbed by the suspended or streambank material contacted. Changes in soluble P concentration during stream flow may occur with the entry of subsurface runoff having a low soluble P con- centration or surface runoff having a high concentration, respectively.

The above processes assume that sufficient desorbable P is present on the sediment for the EPC, to be reached and that the rate of desorption or con- tact time is sufficient for equilibrium to occur during runoff. If the sediment concentration of stream flow is high, then equilibrium may be attained due to rapid P desorption quickly reaching the soluble P concentration in equilibrium with sediment (Kunishi et al., 1972; Schuman et al., 1973; Mc- Coll et al., 1975). The input of sediment from heavily P-fertilized soils may increase the soluble P concentration of stream flow dramatically (Taylor and Kunishi, 1971). If, however, the sediment concentration of stream flow is low, the attainment of the EPC, will be limited by the capacity of the desorbable P pool of the sediment contacted. In this case, the reaction mainly occurs with streambank and bottom material that the stream con- tacts on its way to the watershed outlet. Streambank material is usually P deficient and has a high P sorption capacity. A decrease in the soluble P concentration during base stream flow, when the sediment concentration

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 31 1

was low, has been observed by Taylor and Kunishi (1971), Gburek and Heald (1974), Johnson et al. (1976), and Sharpley and Syers (1979). Soluble P concentrations of 0.10-0.13 mgfliter of runoff from fertilized fields were reduced to 0.009 mg/liter by sorption during movement downstream (Kunishi et al., 1972).

A linear inverse relationship between soluble P concentration and logarithm of sediment concentration of runoff from cropped and grassed watersheds has been observed (Holt et al., 1973; Burwell et al., 1975; Nielson and Mackenzie, 1977; Sharpley et al., 1981~). Sharpley et al. (1981~) reported that the slope of the relationship for several watersheds was related to the P sorption capacities of suspended particulate material in runoff from the watersheds.

It is apparent, therefore, that changes in bioavailability of P can occur be- tween the point where it enters runoff flow and where it enters a lake or im- poundment. Although some unavailable P forms may be converted to available forms in transit downstream, data reported suggest that the predominant reac- tion causes available forms to be converted to unavailable forms. Consequent- ly, the extent to which transformations between soluble and particulate P occur during stream flow must be considered in terms of modeling P movement and the potential bioavailability of P in surface waters.

IV. IMPACT OF PHOSPHORUS ON THE AQUATIC ENVIRONMENT

A close relationship between the total P concentration of lake water and the average algal standing crop in a wide variety of lakes has been observed (Dillon and Rigler, 1974, 1975; Vollenweider, 1975; Schindler, 1977). The occurrence of algal blooms, dissolved oxygen depletions, and fish kills in Horseshoe Lake, Wisconsin, was partially attributed to high P inputs from agricultural and natural drainage by Peterson et al. (1973). In contrast, however, for the economic production of fish, ponds usually require the continuous addition of fertilizer (US Department of Agriculture, 1971). The high production of aquatic plants results in greater fish poundage, due to an increased worm, insect larvae, and other aquatic animal community feeding on the plants. In addition, the high fertility can reduce light penetra- tion. The U.S. Department of Agriculture (1971) recommends that 112 kg of 8-8-2 fertilizer per surface hectare of water be used for the first 3-5 years, with a subsequent annual application of 45 kg/ha superphosphate.

The forms and amounts of P in lake systems are a function of the input of P from external sources, its output from the lake, and the interchange of P among the various sediment and water components. The interaction be- tween soluble and particulate P forms is controlled by chemical,

312 A. N. SHARPLEY AND R. G. MENZEL

biochemical, and physical processes. Although soluble P is immediately available for algal uptake, particulate P may provide a long-term source of available P to algae growth through desorption to the surrounding lake water (Bjork, 1972; Larsen et al., 1975; Cooke et al., 1977). Thus, the pro- cesses controlling the bioavailability of particulate P must be considered in designing programs to control accelerated eutrophication. Soluble and par- ticulate P may be removed from the biotic zone by the natural processes of phytoplankton uptake and deposition. The process of accelerated eutrophication has been temporarily reversed in several eutrophic and hypereutrophic lakes by the inactivation of biologically available P with the addition of alum (Peterson et al., 1973; Cooke et al., 1978).

A. SOLUBLE PHOSPHORUS

The most available form of P to algae in the aquatic environment is solu- ble P (Vollenweider, 1968; Bartsch, 1969). Walton and Lee (1972) reported that soluble P was essentially 100% available, using algal assay procedures and a variety of waters. A number of investigators, however, have found that soluble P as measured by the molybdate method (Murphy and Riley, 1962) is not completely available to support algal growth (Rigler, 1968; Lean, 1973a,b: Dick and Tabatabai, 1977; Stainton, 1980). This results from a possible reduction in condensed phosphates, hydrolysis of organic P compounds, and reaction with arsenate during analysis, all of which will contribute to an overestimation of the true soluble P concentration. This discrepancy is relatively great for waters of low P concentration, such as are normally found in lakes, while the percentage error is much lower with con- centrations found in streams, rivers, or wastewater discharges. Lee et al. (1979) suggested that from a lake management point of view, the discrepan- cy at low soluble P concentrations is of no major consequence as P control programs must be directed at sources of high concentrations.

Boyd and Musig (1981) observed that planktonic communities in samples of water from fish ponds absorbed an average of 41 070 of 0.30 mg/liter addi- tions of soluble P within 24 hr. Over a longer period of time (2 weeks), these concentrations declined to 10% of that originally present due to the added removal of P by sediment.

B. PARTICULATE PHOSPHORUS

In oligotrophic and sometimes in eutrophic waters where soluble P con- centrations are depleted by vigorous algal growth, concentrations may be as low as 0.001 mg/liter (McColl, 1972). Under these conditions, P may be desorbed from the suspended or deposited sediment material. In fact, Ban- nerman et al. (1975) calculated that approximately 10% of the external P

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 313

loading of Lake Erie (1.3 x lo7 kg/ha) resulted from the desorption of P from lake sediments. Furthermore, several studies have reported that par- ticulate P can support biological growth, even though the soluble P concen- tration is low (Healy and McColl, 1974; Golterman, 1977; Allan and Williams, 1978). Consequently, the capacity of sediment entering lakes to supply or remove P is important in an evaluation of the importance of P in the aquatic environment.

The processes involved in the sorption and desorption of P by sediment material in a lake are analogous to those occurring in stream flow discussed earlier. The direction of P movement will be governed by the soluble P concen- tration of the lake water and the desorbable P content or EPC, of the sediment material. The rate and extent of P interchange between sediment P and the sur- rounding lake water is controlled by the forms of P contained in the sediment and soluble P concentration of the interstitial water. The forms of P contained in the sediment have been reviewed extensively by Syers et al. (1973b). The potentially mobile forms are P sorbed on hydrous Fe and Al oxides and CaCO,.

In addition to the chemical mobility of sediment material, its physical mobili- ty will also affect the interchange between particulate and soluble P, and subse- quently its bioavailability. The physical mobility of sediment entering a lake will be a function of its texture and lake water temperature and turbulence. The relative densities and temperatures of the inflow and lake water will determine whether sediment enters the surface or bottom waters of the lake. As it enters the lake its turbidity may reduce the depth of the photic zone. Coarse-textured sediments will settle rapidly and be available to algae in the photic zone for short periods only. In contrast, fine-textured sediments will remain in the photic zone for a longer period of time. The bioavailability of this sediment will be further increased by the fact that it will be enriched in P compared to coarser material. Removal of fine-textured sediments from the photic zone may be enhanced by bioflocculation in the presence of certain algae (Avnimelech and Menzel, 1984).

The mobilty of particulate P can increase if the sediment settles from an oxic photic zone into a deoxygenated hypolimnion. The desorbed P can then be redistributed during periods of lake turnover. In a study of the P dynamics of two shallow hypereutrophic lakes in Indiana, Theis and Mc- Cabe (1978) found that the soluble P concentration of lake water was reduc- ed by sorption during oxic periods and increased by desorption during anaerobic periods. The increased mobility of particulate P under anaerobic compared to aerobic conditions is attributed to a reduction of Fe(II1) to Fe(I1) (Li et al., 1972; Syers et al., 1973b; Patrick and Khalid, 1974).

c. BIOAVAILABILITY OF PHOSPHORUS

Due to the importance of sediment as the major source of P entering the aquatic environment from agricultural land and its ability to sustain algal

3 14 A. N. SHARPLEY AND R. G . MENZEL

growth, several methods to estimate the bioavailability of this P source have been proposed. The availability of P to algae can be determined by an algal culture test (EPA, 1971). However, more rapid chemical extraction pro- cedures, which simulate removal of P by algae, have been proposed for the routine determination of particulate P bioavailability (Dorich et al., 1985). Chemical extractants that have been used to measure the bioavailability of particulate P are NaOH (Sagher et al., 1975, Golterman, 1976; Cowan and Lee, 1976; Armstrong et al., 1979; Logan et al., 1979); NH,F (Porcella et al., 1970; Dorich et al., 1980); anion exchange resins (Wildung and Schmidt, 1973; Cowan and Lee, 1976; Armstrong et al., 1979; Huettl et al., 1979); and citrate-dithionite-bicarbonate (CDB) (Logan et al., 1979). It is suggested that the weaker extractants and short-term resin extractions represent P that could be utilized by algae in the photic zone of lakes under aerobic conditions. In contrast, the more severe extractants (CDB) repre- sent P that might become available under reducing conditions found in the anoxic hypolimnion of stratified lakes.

Caution must be exercised, however, in relating P bioavailability of sedi- ment material determined by these chemical extractions and the potential of the sediment to increase algal growth (Lee et al., 1979; Sonzogni et al., 1982). In turbid stratified lakes the surface photic zone may be relatively thin compared to the mixed layers above or below the thermocline. In addi- tion, suspended sediment material often contains large amounts of silt-sized aggregates of clay, which will settle more rapidly from the photic zone than smaller particles, possibly reducing the actual availability of particulate P. Consequently, bioassays may produce erroneous estimates of available par- ticulate P unless the physiochemical properties of the waterbody and sedi- ment are considered in determining the appropriate bioassay to be used.

Although these above chemical extraction procedures have identified which particulate P fractions can be utilized by algae, there is no evidence that all of the chemically extracted P is algal-available. Thus, Hegemann et al. (1983) suggested that a quantitative assessment of algal-available par- ticulate P will depend upon the development of long-term (> 100 day) algal assay procedures. The bioavailability of P attached to suspended sediment transported in tributaries to lakes and of sediment deposited in lakes is sum- marized in Table IV for several studies. It is evident that a large variability in the bioavailability of sediment P exists, which reflects the dynamic nature of the physiochemical processes governing the transport, P mobility, and deposition of eroded soil material. Phosphorus associated with suspended sediment can be considered to be of short-term bioavailability due to sedimentation from the biotic zone.

In contrast, P associated with deposited sediments is potentially bioavailable for a much longer period of time. Wildung et al. (1974) reported that the P content of the sediment in several lakes in Oregon was

Table N

Percentage Bioavailability of Sediment P Transported in Several Lake Tributaries Draining Agricultural Watersheds and in Deposited Lake Sediments

Bioavailabilitf Total P Reference Location Description Procedure (Qo) Wkg)

Suspended sediment in tributaries Dorich et al. (1980) Indiana Agricultural Bioassay 21 0.2-0.7

NH, F 9 NaOH 8 HCl 4

NaOH 4-38 CDB* 9-27

De Pinto el al. (1981) Great Lakes Agricultural Bioassay 0-4 0.5-1.2

Logan et al. (1979) Lake Erie Agricultural NaOH 14-42 0.5-1.2 CDB 29-56

Deposited sediment Allan and Williams (1978) Central Canada Prairie Lakes CDB 14-37 0.5-1.3 Bannerman et al. (1975) Lake Ontario Postglacial

Glacial Basin

Carigan and Kalff (1980) Quebec C. L. Memphremagog S. L. Memphremagog Riviere-du-sud

Klapwijk et al. (1982) Netherlands Rijuland water Sagher et al. (1975) Wisconsin Calcareous

Noncalcareous Williams et 01. (1980) Great Lakes Bottom sediments

NaOH NaOH NaOH Resin Resin Resin Bioassay NaOH NaOH

NaOH Resin

N T A ~

30-60 1 .O-1.5 2-8 0.9-1.0 13-18 0.8-0.9 8 0.8-1.2 25 19

0-41 0.4-4.8 60-95 0.6-3.9 80-85 30 0.4-1.4 27 21

'Tercentage total particulate P bioavailable. bCDB and NTA represent citrate-dithionate-bicarbonate and 0.01 Mneutralized nitdoacetic acid extractable P, respectively.

316 A. N. SHARPLEY AND R. G. MENZEL

directly related to the biological productivity of surface waters and served as a significant source of P to these waters, supporting increased biological growth. Carignan and Kalff (1980) found that submerged macrophytes depended overwhelmingly on sediments for their P supply. Even under hypereutrophic lake conditions, sediments contributed the major propor- tion (72010) of P utilized during growth. It has been suggested, moreover, that these aquatic plants may supply P to overlying waters by excretion dur- ing growth and upon senescence (Carignan and Kalff, 1980).

The water renewal time of a lake plays an important role in the dynamics and extent of P exchanges in a lake. With a short residence time, outflow of water from a lake can be a more important route for P removal than sedimentation. When the residence time of a lake exceeds a few months, most of the P inflow is retained in the lake sediments. Because of this pro- cess, impoundments and small lakes have been used as efficient traps (especially for particulate P) to improve downstream water quality (Rausch and Schreiber, 1977). It is apparent, however, that the amounts of P stored in lakes can build up to unacceptable levels, resulting in a permanent deterioration in water quality. In fact, a reduction in the external load of P upon the highly eutrophic Lake Trummen in Sweden did not bring about the desired improvement in water quality until the upper layers of the P-rich sediment were removed (Bjork, 1972).

D. ARTIFICIAL REMOVAL OF PHOSPHORUS FROM LAKES

In order to control or reduce the increased biological productivity of lakes and impoundments, the inputs of P must first be reduced. The diver- sion of P inputs, however, does not always bring about a prompt and suffi- cient reduction in lake water concentration, due to internal recycling from P-rich sediments (Larsen et al., 1975; Cooke et al., 1977). A reduction in the P concentration of lake water and the inactivation of the recycling mechanisms may be brought about by chemical amendments (Jernelov, 1970; Peterson et al., 1973; Cooke et al., 1978; Kennedy, 1978). Several points need to be considered in the restoration of lake water quality by P precipitation and inactivation. These include the chemicals used, dosage, ef- fect of the additives on benthic fauna, and method and time of application. The most commonly used chemicals are aluminum sulfate and sodium aluminate, due to the stability of flocculated Al hydroxides with redox changes. The removal of P is brought about by precipitation of AlPO,, by coagulation or entrapment of P-containing particulates, or by sorption of P on the surfaces of A1 hydroxide polymers (Recht and Ghassemi, 1970; Eisenreich et al., 1977).

The maximum dosage of A12(S0,)3 for the long-term control of P cycling may be determined by Al,(S04)3 addition to lake water samples until the

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 317

dissolved Al concentration reaches 0.050 mg AlAiter (Kennedy, 1978), a concentration Everhart and Freeman (1973) found to be nontoxic to fish. Very little direct laboratory or field evidence on the effect of Al on the aquatic biota exists, although several studies have shown no apparent effect on fish (Kennedy and Cooke, 1974; Bandow, 1974; Sanville et al., 1976) or benthic invertebrates (Narf, 1978) following full-scale lake treatments.

A predetermined amount of A12(S04), is applied as a slurry from the lake surface if P removal from the epilimnion is required. If control of P release from sediments is required then application to the hypolimnion is necessary. As d2(so4)3 removes dissolved organic P inefficiently (Browman et al., 1973; Eisenreich et al., 1977), applications should be made in early spring when the major proportion of P in lake water is inorganic (Browman et al., 1977; Eisenreich et af., 1977). The continued presence of organic P may be significant, as Heath and Cooke (1975) observed that certain nuisance blue- green algae can produce a phosphatase enzyme under P-limiting conditions, that is capable of mineralizing organic to inorganic P at rates sufficient to support algal blooms. Application time will not be critical for treatment of P desorption from lake sediments. However, the relative importance of lake sediments as a P source should be assessed prior to Al,(SO4), application. For example, lakes receiving substantial inputs of clay in addition to P may contain sediments with high sorption capacities for P.

The application of A12(S04), to just below the surface of Horseshoe Lake, Wisconsin, resulted in a significant decrease in the P content of both the epilimnion and hypolimnion (Peterson et al., 1973). Prior to application, the lake had experienced algal blooms and fish kills which were partially at- tributed to agricultural inputs of P. Born (1979) observed that although hypolimnion P increased slightly each year after application, it never reach- ed pretreatment levels, thus giving approximately 8 years of control. The hypolimnetic application of Al,(S04), to the eutrophic West Twin Lake, Ohio, resulted in an 88% reduction in total P concentration of the lake water (Kennedy, 1978). Continued water quality monitoring by Kennedy (1978) indicated that the layer of Al(OH), deposited on the sediments re- duced P release to overlying waters by 98%. Three years later the lake was mesotrophic.

V. CONCLUSIONS

Fertilizer P use presents no direct problem to the terrestrial environment. The use of P fertilizer is essential to maintain adequate crop production for an ever-increasing population. Its application can reduce the nutrient enrichment of surface waters by establishing an increased vegetative cover

318 A. N. SHARPLEY AND R. G . MENZEL

on eroding soils. These benefits to the environment must be considered along with a potentially detrimental indirect effect of fertilizer P use. Heavy metal and radionuclide contaminants not removed during fertilizer manufacture may xcumulate in the soil. No threat to human health has been reported at the present time, although the continued application of im- pure fertilizer materials to acid soils may lead to problems with crops susceptible to contaminant uptake, especially that of cadmium.

Most of the public and scientific attention regarding environmental ef- fects of P fertilizer has been focused on the aquatic environment due to the role of P in increasing the biological productivity of lakes and impound- ments. Considerable research has been conducted to quantify the losses of soil and fertilizer P from various land management practices. However, we are still unable to relate P inputs to a lake or impoundment to a quantitative description of water quality. Furthermore, the effect of P concentration on algal growth receives continued attention, while little information is available on how lake macrophytes are affected, even though macrophytes present a more serious economic problem than algae in many lakes.

Research should be directed towards improving the partitioning models for soluble and particulate P transport in runoff and in lakes and impound- ments. This should focus on the mechanisms of exchange between desor- bable or labile P and solution and methods to routinely quantify the amounts of desorbable or bioavailable P on various materials. With the ac- cumulation of P at the soil surface under conservation tillage practices, ex- isting soil P test procedures may need to be reevaluated. This may include changes in soil sampling frequency, timing, and depth, in addition to the use of chemical extractants capable of removing easily mineralizable organic P from the surface plant residue built up.

As the crop canopy can contribute a major proportion of the soluble P transported in runoff, surface soil and plants must be considered as a con- tinuum and the pool of desorbable P in both soil and plant material deter- mined. The measured size of this pool will depend upon the experimental conditions of analysis, therefore extraction mediums, so1ution:soil ratios, and contact times relevant to either the terrestrial or aquatic environments must be used. In the case of the aquatic environment, the fact that the desorbed P can be continuously removed from the system by algal growth must be considered.

In the light of research on the kinetics of P exchange between desorbable P and solution in the terrestrial and aquatic environments and during transport from the terrestrial environment, more accurate and widely ap- plicable models simulating P transport from watersheds can be expected. This information should be used to improve the prediction of both the amounts and forms of P transported into lakes and impoundments. These models can then be used as tools to aid management decisions to reduce P

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 319

loss in runoff and at the same time to increase crop yields to maintain ade- quate food production for an increasing population. In addition, these models may also identify areas of further research.

ACKNOWLEDGMENT

The authors wish to acknowledge the pioneering phosphorus and water quality research of Dr. J. C. Ryden, who died suddenly May 3, 1986, in London, England.

REFERENCES

Adriano, D. C., Novak, L. T., Erickson, A. E., Wolcott, A. R., and Ellis, G. E. 1975. J. En-

Ahuja, L. R., Sharpley, A. N., Yamamoto, M., and Menzel, R. G. 1981. Water Resour. Res.

Allan, R. J., and Williams, J. D. H. 1978. J. Environ. Qual. 7, 990-106. Andersson, A,, and Mahlin, M. 1981. Swed. J. Agric. Res. 11, 3-10. Andersson, A., and Nielson, K. 0. 1974. Ambio 3, 198-200. Andraski, B. J., Muller, D. H., and Daniels, T. C. 1985. Soil Sci. SOC. Am. J. 49, 1523-1527. Andrews, K. C., Muller, F. B., McSweeney, G., and Seager, R. H. 1979. Proc. Tech. Conf., N.

Z . Fertilizer ManuJ. Res. Assoc., 17th, Aukland pp. 321-348. Armstrong, D. E., Perry, J. R., and Flatness, D. 1979. Final Rep. Wis. Water Res. Cent.,

Madison. Avnimelech, Y., and Menzel, R. G. 1984. J. Soil Water Conserv. 39, 200-203. Bache, B. W., and Ireland, C. 1980. J. Soil Sci. 31, 297-306. Bahnick, D. A. 1977. J. Environ. Qual. 6 , 217-222. Baker, D. E. 1964. SoilSci. SOC. Am. Proc. 28, 511-518. Baker, J. L., Campbell, K. L., Johnson, H. P., and Hanway, J. J. 1975. J. Environ. Qual. 4,

Ballaux, J. C., and Peaslee, D. E. 1975. Soil Sci. SOC. Am. Proc. 39, 275-178. Bandow, F. 1974. Minn. Dep. Nut. Resour. Div. Fish Wildlije Invest. Rep. 326. Bannerman, R. T., Armstrong, D. E., Harris, R. F., and Holdren, C. C. 1975. Ecol. Res. Ser.

Barber. S. A. 1979. Commun. Soil Sci. Plant Anal. 10, 1459-1468. Barrow, N. J. 1978. J. SoilSci. 29, 447-462. Barrow, N. J. 1979. J. SoilSci. 30, 259-270. Barrow, N. J., and Shaw, T. C. 1975. Soil Sci. 119, 167-77. Barrow, N. J., and Shaw, T. C. 1979. J. Soil Sci. 30, 67-76. Barrows, H. L. 1966. J. Soil Water Conserv. 21, 211-216. Barisas, S. G., Baker, J. L., Johnson, H. P., and Laflen. J. M. 1978. Trans. Am. SOC. Agric.

Bartsch, A. F. 1969. Provisional algal assay procedures. Joint Ind. Gov. Tusk Force

Baule, H. 1973. Potash Rev. 21/22. Bjork, S . 1972. Ambio 1, 153-165. Bolton. E. F., Aylesworth, J. W., and Hare, F. R. 1970. Can. J. Soil Sci. 50. 275-279. Born, S. M. 1979. Environ. Manage. 3, 145-153. Bowen. H. J. M. 1966. “Trace Elements in Biochemistry.” Academic Press, London.

viron. Qual. 4, 242-248.

17, %9-974.

406-412.

EPA-660/3-75/-066, pp. 61.

Eng. 21, 893-897.

Eutrophicat., New York.

320 A. N. SHARPLEY AND R. G. MENZEL

Boyd, C. E., and Musig, Y. 1981. Aquaculture 22, 165-173. Bromfield, S. M., and Jones, 0. L. 1972. Aust. J. Agric. Res. 23, 811-824. Browman, M. G., Harris, R. F., and Armstrong, D. E. 1973. Renewal by treatment with

Browman, M. G., Harris, R. F., and Armstrong, D. E. 1977. Tech. Rep. No. 77-05, Water

Burwell, R. E., and Kramer, C. A. 1983. J. Soil Water Conserv. 38, 315-319. Burwell, R. E., Schuman, G. E., Piest, R. F., Spomer, R. G., and McCalla, T. M. 1974. Water

Burwell, R. E., Timmons, D. R., and Holt, R. F. 1975. Soil Sci. SOC. Am. Proc. 39, 523-528. Burwell, R. E., Schuman, G. E., Heinemann, H. G., and Spomer, R. G. 1977. J. Soil Water

Campbell, D. A. 1950. Trans. Znt. Congr. SoilSci. 1, 327-331. Carignan, R., and Kalff, J . 1980. Science Un, 987-989. Caro, J. H. 1964. In “Superphosphate: Its History, Chemistry, and Manufacture,” (J. 0. Hardesty

Carter, D. L., Bondurant, J. A., and Robbins, C. W. 1971. Soil Sci. SOC. Am. Proc. 35,

C.A.S.T. (Council for Agricultural Science and Technology). 1976. EPA430/9-76-013.

Cooke, G. D., McComas, M. R., Walker, D. W., and Kennedy, R. H. 1977. Hydrobiologia

Cooke, G. D., Heath, R. T., Kennedy, R. H., and McComas, M. R. 1978. EPA-600/3-78-003.

Cowen, W. F., and Lee, G. F. 1973. Environ. Sci. Technol. 7, 853-854. Cowen, W. F., and Lee, G. F. 1976. Ecolog. Res. Ser.. EPA-600/3-76-094a. USEPA,

DePinto, J. V., Young, T. C., and Martin, S. C. 1981. J. Great Lakes Res. 7, 557-568. Dick, W. A,, and Tabatabai, M. A. 1977. J. Environ. Quai. 6, 82-85. Dillon, P. J., and Rigler, F. H. 1974. Limnol. Oceanogr. 19, 767-773. Dillon, P. J., and Rigler, F. H. 1975. J. Fish. Res. Board Can. 32, 1519-1531. Donigian, A. G., Jr., Beyerlein, D. B., Davis, H. H., and Crawford, N. H. 1977. Report No.

Dorich, R. A., Nelson, D. W., and Sommers, L. E. 1980. J. Environ. Quai. 9, 557-563. Dorich, R. A., Nelson, D. W., and Sommers, L. E. 1985. J. Environ. Quai. 14, 4oo-405. Duffy, P. D., Schreiber, J. D., McClurkin, D. C., and McDowell, L. L. 1978. J. Environ. Quai.

Duxbury, J. M., and Peverly, J. H. 1978. J. Environ. Quai. 7 , 566-570. Eisenreich, S. J., Armstrong, D. E., and Harris, R. F. 1977. Tech. Rep., Water Resour. Center,

Elrashidi, M. A., and Larsen, S . 1978. Plant Soil 50, 585-5%. EPA. 1971. Assay procedure-Bottle Test. National Eutrophication Res. Program, Pacific

Evans, R. L., and Jurinak, J. J. 1976. Soil Sci. 121, 205-21 1. Everhart, W. H., and Freeman, R. A. 1973. Toxic effects of aqueous aluminum to rainbow

Forster, D. L., Logan, T. J., Taksich, S. M., and Adams, J. R. 1985. J. Soil Water Conserv. 40,

Fox, R. L., and Kamprath, E. J. 1971. Soil Sci. Am. Proc. 35, 154-156. Fukely, G. 1978. Commun. Soil Sci. Plant Anal. 9, 851-863. Garcia-Miragaya, J., and Page, A. L. 1978. Water, Air Soil Pollut. 10, 259-299. Gardner, B. R., and Preston Jones, J . 1973. Commun. Soil Sci. Plant Anal. 4, 83-93.

aluminum hydroxide. Report to Wisconsin Dept. Nat. Resour., Madison.

Resour. Cent., Univ. of Wisconsin, Madison.

Resour. Res. 10, 359-364.

Conserv. 32, 266-230.

and A. V. Slack, eds.), pp. 272-305. U.S. Dept. Agric. and TVA, Washington, D. C.

33 1-335.

USEPA, Washington, D. C.

56, 129-135.

USEPA, Washington, D. C.

Washington, D. C.

EPA 600/3-77-098, USEPA, Athens, Georgia.

7, 45-50.

No. 77-02, University of Wisconsin, Madison.

North West Laboratory, Corvdis, Oregon.

trout. EPA-R3-73-01l.b.

136-141.

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 32 1

Gburek, W. J., and Broyan, J. G. 1974. Proc. Agric. Waste Manage. Conf. pp. 39-50. Gburek, W. J., and Heald, W. R. 1974. Water Resour. Res. 10, 113-118. Golterman, H. L. 1976. In “Interactions between Sediments and Freshwater” (H. K. Golter-

man, ed.). Junk and Pudoc, The Hague. Golterman, H. L. 1977. In “Proc. Workshop on Fluvial Transport of Sediment Associated

Nutrients and Contaminants” (H. Shear, ed.). IJC-PLUARG, Windsor, Ontario, Canada. Goodroad, L. L., and Caldwell, A. C. 1979. J. Environ. Qual. 8, 493-4%. Gosz, J. R. Likens, G. E., and Borman, F. H. 1973. Ecology 53, 769-784. Gotoh, S., and Patrick, W. H., Jr. 1974. Soil Sci. Soc. Am. Proc. 38, 66-71. Green, D. B., Logan, T. J., and Smeck, N. E. 1978. J. Environ. Qual. 7, 208-212. Hanway, J. J., and Laflen, J. M. 1974. J. Environ. Qual. 3, 351-356. Healy, W. B., and McColl. R. H. S. 1974. N.Z. J. Sci. 17,409420. Heath, R. T., and Cooke, G. D. 1975. Verh. Int. Ver. Limnol. 19, 959-965. Hegemann, D. A., Johnson, A. H., and Keenan, J. D. 1983. J. Environ. Qual. 12, 12-16. Holt, R. F., Johnson, H. P., and McDowell, L. L. 1973. Proc. Nut. Conserv. Tillage Corlf.,

Hope, G. D., and Syers, J. K. 1976. J. Soil Sci. 27, 301-306. Hortensteine, C. C., and Forbes, R. B. 1972. J. Environ. Qual. 1, 446-449. Hostetler, P. B., and Carrels, R. M. 1%2. Econ. Geol. 57, 137-167. Huettl, P. J., Wendt, R. C., and Corey, R. B. 1979. J. Environ. Qual. 8, 541-548. Jarvis, S. C., and Jones, L. H. P. 1980. J. SoilSci. 31, 469479. Jernelov, A. 1970. Vatten 26, 262-272. Johnson, A. H., Boddin, D. R., Goyette, E. A., and Heges, A. N. 1976. J. Environ. Qual. 5,

Jones, L. A., Smeck, N. E.. and Wilding, L. P. 1977. J. Environ. Qual. 6, 2%-302. Kennedy, R. H. 1978. Ph.D. dissertation, Kent State Univ., Diss. Abstr. 2143-B. Kennedy, R. H., and Cooke, G. D. 1974. Conf. Lake Protect. Manage., Madison, Wisconsin. Khalid, R. A., Patrick, W. H., Jr, and Delaune, R. D. 1977. Soil Sci. Soc. Am. J. 41, 305-310. Klages, M. G., and Hsieh, Y. P. 1975. .I. Environ. Qual. 4, 68-73. Klapwijk, S. P., Kroon, J. M. W., and Meijer, M. L. 1982. Hydrobiologia 92, 491-500. Kleusner, J. 1972. Ph.D. thesis, Univ. of Wisconsin, Madison, Diss. Abstr. 72-22, 100. Knoblauch, H. C., Kolosky, L., and Brill, G. D. 1942. Soil Sci. 53, 369-389. Kunishi, H. M., Taylor, A. W., Heald, W. R., Gburek, W. J., and Weaver, R. N. 1972. J.

Agric. Food Chem. 20, 900-905. Lagerwerff, J. V. 1971. Soil Sci. 111, 129-133. Langdale, G. W., Perkins, H. F., Barnett, A. P., Reardon, J. C., and Wilson, R. L., Jr. 1983. J.

Soil Water Conserv. 38, 297-301. Langdale, G. W., Leonard, R. A,, and Thomas, A. W. 1985. J. Soil Water Conserv. 40,

Larsen, D. P., Maleg, K. W., Schultz, D. W., and Brice, R. M. 1975. Verh. Int. Ver. Limnol. 19,

Lean, D. R. S . 1973a. Science 179, 678-679. Lean, D. R. S. 1973b. J. Fkh. Res. Board Can. 30, 1525-1536. Lee, G. F., 1973. Water Ra. 7, 111-128. Lee, G. F., Rast, W., and Jones, R. A. 1978. Environ. Sci. Technol. 12, 900-908. Lee, G. F., Jones, R. A., and Rast, W. 1979. In “Phosphorus Management Strategies for Lakes”

(R. C. Leohr, C. S. Martin, and W. Rat. eds.), pp. 259-308. Ann Arbor Sci. Publ., Ann Arbor, Michigan.

Li, W. C., Armstrong, D. E., Williams, J. D. H., Harris, F. F., and Syers, J. K. 1972. Soil Sci. Soc. Am. Proc. 36, 279-285.

Lisk, D. F. 1972. Adv. Agron. 24, 268-325. Logan, T. J., and Adams, J. R. 1981. LEWMS Tech. Rep., U.S. Army Corps Eng., Buffalo,

March, Des Moines pp. 141-156.

148-157.

157-160.

884-892.

New York.

322 A. N. SHARPLEY AND R. G. MENZEL

Logan, T. J., Oloya, T. O., and Yaksich, S. M. 1979. J. Great Lakes Res. 5, 112-123. McCallister, D. L., and Logan, T. J. 1978. J. Environ. Qual. 7, 87-92. McCoU, R. H. S. 1972. N.Z. J. Mar. Freshwater Res. 6, 399447. McCoU, R. H. S. 1974. N.Z. J. Mar. Freshwater Res. 8, 375-388. McColl, R. H. S., White, E., and Waugh, J. R. 1975. N.Z. J. Sci. 18, 67-84. McColl, R. H. S., White, E., and Gibson, A. R. 1977. N.Z. J. Mar. Freshwater Res. 11,

McDowell, L. L., and McGregor, K. C. 1984. Soil Till. Res. 4, 79-91. McDowell, L. L., Schreiber, J. D., and Pionke, H. B. 1980. In “CREAMS-A Field Scale

Model for Chemical Runoff and Erosion from Agricultural Management Systems. Vol. 111. Supporting Documentation” (W. Knisel ed.), pp. 509-533. USDA, Conservation Res. Rep. No. 26.

Marsden, E. 1964. In “The Natural Radiation Environment” (J. A. S. Adams and W. M. Lowder, eds.), pp. 807-824. Univ. of Chicago Press, Chicago.

Massey, H. F., and Jackson, M. L. 1952. SoilSci. SOC. Am. Proc. 16, 353-356. Mays, P. A., and Mortvedt, J. J. 1986. J. Environ. Qual. 15, 78-81. Menzel, R. 0. 1968. Agric. Food Chem. 16, 231-234. Menzel, R. G. 1980. In “CREAMS-A Field Scale Model for Chemicals Runoff and Erosion

from Agricultural Management Systems. Vol. 111. Supporting Documentation” (W. Knisel ed.), pp. 486-492. USDA, Conservation Res. Rep. No. 26.

Menzel, R. G., Rhoades, E. D., Olness, A. E., and Smith, S. J. 1978. J. Environ. Qual. 7, 4 0 1 4 .

Miller, M. H. 1979. J. Environ. Qual. 8, 42-48. Moldenhauer, W. C., Langdale, G. W., Fryre, W., McCool, D. K., Papendick, R. I., Smika,

Mortvedt, J. J. 1987. J. Environ. Qual. 16, 137-142. Mortvedt, J. J., and Giordano, P. M. 1977. Hanford Life Sci. Symp., 15th, 29 Sept-10 Oct,

1975, Richland, Washington. ERDA Rep. Conf. 750929. Natl. Tech. Inf. Serv., Springfield, VA.

729-744.

D. E., and Fryrear, D. W. 1983. J. Soil Water Conserv. 38, 144-151.

Mortvedt, J. J., Mays, D. A., and Osborn, D. 1981. J. Environ. Qual. 10, 193-197. Muir, J., Olsen, E. C., and Olsen, R. A. 1973. J. Environ. Qual. 2, 466-470. Mulla, D. A., Page, A. L., and Ganje, T. J. 1980. J. Environ. Qual. 9, 408-412. Murad, E., and Fischer, W. R. 1978. Geoderma 21, 133-145. Murphy, J., and Riley, J. P. 1962. Anal. Chim. Acta 27, 31-36. Murphy, T. J., and Doskey, P. V. 1975. Inputs of phosphorus from precipitation. USEPA Rep.

Narf, R. P. 1978. An evaluation of past aluminum sulfate lake treatments: Present sediment

Neal, 0. R. 1944. Agron. J. 36, 601-607. Nelson, D. W., Monke, E. J., Bottcher, A. D., and Sommers, L. E. 1979. In “Best Management

Practices for Agriculture and Silviculture” (R. C. Loehr, D. A. Haith, M. F. Walter, and C. S. Martin, eds.), pp. 491-505. Ann Arbor Sci. Publ., Ann Arbor, Michigan.

600/3-75-005, Duluth, Mn.

aluminum concentrations and benthic insect renewal. Wisc. Rep. Natl. Res., Madison.

Nicholaichuk, W., and Read. D. W. L. 1978. J. Environ. Qual. 7 , 542-544. Nielson, G . H., and Mackenzie, A. F. 1977. Can. J. Soil Sci. 57, 465-475. Oloya, T. O., and Logan, T. J. 1980. J. Environ. Qual. 9, 526-531. Ozanne, P. G., Kirton, D. J., and Shaw, T. C. 1961. Aust. J. Agric. Res. 12, 409-423. Parfitt, R. L. 1978. Adv. Agron. 30, 1-50. Patrick, W. H., and Khalid, R. A. 1974. Science 186, 53-55. Peterson, J. O., Wall, J. J. Wirth, T. H., and Born, S. M. 1973. Tech. Bull. No. 62, Wisc. Dept.

of Natl. Resour., Madison, Wisconsin.

THE IMPACT OF SOIL AND FERTILIZER PHOSPHORUS 323

Ponnamperuma, F. N. 1972. SoilSci. SOC. Am. J. 41, 305-310. Porcella, D. B., Kumagai, J. S., and Middlebrooks, E. J. 1970. J. Sanit. Eng. Div., Proc. Am.

Porter, K. S . 1975. “Nitrogen and Phosphorus, Food Production, Waste and the Environment.”

Randall, G. W. 1980. Proc. Annu. Fertilizer Agric. Chem. Dealers Conf., 32nd, Iowa State

Rausch, D. L., and Schreiber, J. D. 1977. Trans. Am. SOC. Agric. Eng. 20, 281-290. Recht, H. L., and Ghassemi, M. 1970. Water Pollut. Control Res. Serv. 17010 EKI. Reedy, G. Y., McLean, E. O., Hop, G. D., and Logan, T. J. 1978. J. Environ. Qual. 7,5044. Reuss, J. O., Dooley, H. L., and Griffin, W. 1978. J. Environ. Qual. 7 , 128-133. Rhoton, F. E., Smeck, N. E., and Wildung, L. P. 1979. J. Environ. Qual. 8, 547-550. Rigler, F. H. 1%8. Limnol. Oceanogr. 13, 7-13. Rogers, M. T. 1941. Soil Sci. Soc. Am. 6 , 263-271. Romkens, M. J. M., and Nelson, D. W. 1974. J. Environ. Qual. 3, 10-13. Romkens, M. J. M., Nelson, D. W., and Mannering, J. V. 1973. J. Environ. Qual. 2,292-298. Rothbaum, H. P., McGaveston, D. A., Wall, T., Johnson, A. E., and Mattingly, G. E. G.

Ryden, J. C., and Syers, J. K. 1977a. Soil Sci. 123, 353-361. Ryden, J. C., and Syers, J. K. 1977b. J. Soil Sci. 28, 596-609. Ryden, J. C., Syers, J. K., and Harris, R. F. 1972. J. Environ. Qual. 1, 431-434. Ryden, J. C., Syers, J. K., and Harris, R. F. 1973. Adv. Agron. 25, 1-45. Sagher, A., Harris, R. F., and Armstrong, D. E. 1975. Univ. of Wisc. Water Resour. Cent.,

Sanville, W. D., Gahler, A. R., Searcy, J. A., and Powers, C. E. 1976. EPA-600/3-76441. Sawhney, B. L. 1977. J. Environ. Qual. 6 , 86-89. Sawhney, B. L., and Frink, C. R. 1978. Clays Clay Mineral Proc. Conf. 26, 227-229. Sawyer, C. N. 1947. J. N. Eng. Water Works Assoc. 61, 109-127. Schindler, D. W. 1977. Science 195, 260-262. Schindler, D. W., and Nighswander, J. E. 1970. J. Fish. Res. Board Can. 27, 260-262. Schreiber, J. D. 1985. J. Environ. Qual. 14, 256-260. Schreiber, J. D., and McDoweU, L. L. 1985. J. Environ. Qual. 14, 251-256. Schreiber, J. D., Rausch, D. L., and McDowell. L. L. 1977. Trans. Am. Soc. Agric. Eng. 20,

Schroeder, H. A., and Balassa, J. J. 1961. J. Chronic Dis. 14, 236-258. Schroeder, H. A.. Nason, A. P., Tipton, I. H., and Balassa, J. J. 1967. J. Chronic Dis. 20,

Schultz, R. K. 1965. Health Phys. 11, 1317-1324. Schuman, G. E., Spomer, R. G., and Piest, R. F. 1973. Soil Sci. SOC. Am. Proc. 37, 424-427. Sharpley, A. N. 1980. J. Environ. Quul. 9, 521-526. Sharpley, A. N. 1981. J. Environ. Qual. 10, 160-165. Sharpley, A. N. 1985a. Soil Sci. Soc. Am. J. 49. 1010-1015. Sharpley, A. N. 1985b. SoilSci. SOC. Am. J. 49, 1527-1534. Sharpley, A. N., and Syers, J. K. 1979. Water, Air Soil Pollut. 11, 417-428. Sharpley, A. N., Syers, J. K., and Tillman, R. W. 1978. J. Environ. Qual. 7, 455-456. Sharpley, A. N., Ahuja, L. R., and Menzel, R. G. 1981a. J. Environ. Qual. 10, 386-391. Sharpley, A. N., Ahuja, L. R., Yamamoto, M., and Menzel, R. G. 1981b. SoilSci. Soc. Am. J.

Sharpley, A. N., Menzel, R. G., Smith, S. J., Rhoades, E. D., and Olness, A. E. 1981c. J. En-

Sharpley, A. N., Smith, S. J., and Menzel, R. G. 1985a. J. Soil Water Conserv. 40, 283-284. Sharpley, A. N., Smith, S. J., Menzel, R. G., and Westerman, R. L. 1985b. Okla. State Agric.

Soc. Civil Eng., SA4, 91 1-926.

Ann Arbor Sci. Publ., Ann Arbor, Michigan.

Univ., Ames.

1979. J. Soil Sci. JO, 147-153.

Tech. Rep. WIC WRC 75-01.

285-290.

179-210.

45, 493-496.

viron. Qual. 10, 211-215.

Exp. Stat. Tech. Bull. T#162.

324 A. N. SHARPLEY AND R. G. MENZEL

Sibbesen, E. 1981. J. Soil Sci. 32, 67-74. Singer, M. J., and Rust, R. H. 1975. J. Environ. Qual. 4, 307-311. Singh, B. B., and Jones, J. P. 1976. Soil Sci. Soc. Am. J. 40, 389-394. Sonzogni, W. C., Chapra, S. C., Armstrong, D. E., and Logan, T. J. 1982. J. Environ. Qual.

Spalding, R. F., and Sackett, W. M. 1972. Science 175, 629-631. Stainton, M. P. 1980. Can. J. Fish. Aguat. Sci. 37, 472478. Stake, E. 1%8. Schweiz. Z. Hydrol. 30, 353-373. Stoltenberg, N. L., and White, J. L. 1953. Soil Sci. Soc. Am. Proc. 17, 406410. Syers, J. K., Browman, M. B., Smillie, G. W., and Corey, R. B. 1973a. Soil Sci. Soc. Am.

Syers, J. K., Harris, R. F., and Armstrong, D. E. 1973b. J. Environ. Qual. 2, 1-14. Tabatabai, M. A., Burwell, R. E., Ellis, B. G., Keeny, D. R., Logan, T. J., Nelson, D. W.,

Olsen, R. A., Randall, G. W., Timmons, D. R., Verry, E. S., and White, E. M. 1981. Iowa State Univ. Agric. Home Econ. Exp. Stat. Res. Bull. 594, 111-142.

Taylor, A. W., and Kilmer, V. J. 1980. In “The Role of Phosphorus in Agriculture,” pp. 545-557. Am. Soc. Agron. Publ., Madison, Wisconsin.

Taylor, A. W., and Kunishi, H. M. 1971. J. Agric. Food Chem. 19, 827-831. Theis, T. L., and McCabe, P. J. 1978. Water Res. 12, 677-685. Timmons, D. R., and Holt, R. F. 1980. In “CREAMS-A Field Scale Model for Chemicals,

Runoff, and Erosion from Agricultural Management Systems. Vol. 111. Supporting Documentation” (W. G. Knisel ed.), USDA Conserv. Res. Rep. 26, 534-543.

11, 555-563.

Proc. 37, 358-363.

Timmons, D. R., Holt, R. F., and Latterell, J. J. 1970. Water Resour. Res. 6, 1367-1375. Tremearne, T. H., and Jacobs, K. D. 1941. USDA Tech. Bull. 781. U.S. Army Corps of Engineers. 1982. Lake Erie wastewater management study. Final Report.

U.S. Department of Agriculture. 1971. Farmer Bull. 2250. Viets, F. G. 1975. C. R. C. Crit. Rev. Environ. Controlpp. 423453. Vig, A. C., Biswas, C. R., and Sinha, M. K. 1979. Bull. Indian Soc. Soil Sci. 12, 279-292. Vincent, W. F., and Downes, M. T. 1980. Aguat. Bot. 9, 221-235. VoUenweider, R. A. 1%8. Scientific fundamentals of the eutrophication of lakes and flowing

waters with particular reference to nitrogen and phosphorus. Organ. Economic Coop. Develop. Rep. OAS/CSI/68.27, pp. 256. Paris, France.

VoUenweider, R. A. 1975. Schweiz. Z. Hydrol. 37, 53-84. Wall, G. J., and Wilding, L. P. 1976. J. Environ. Qual. 5, 168-173. Walton, C. P., and Lee, G. F. 1972. Verh. Int. Ver. Limnol. 18, 676-684. Wang, W. C. 1974. Water Res. Bull. 10, 662-671. Wells, C., Wingham, B., and Lieth, H. 1972. J. Elisha Mitchell Sci. Soc. 88, 66-78. Wells, K. L. 1985. Solutions 29, 3445. Wendt, R. C., and Alberts, E. E. 1984. J. Environ. Qual. 13, 613-618. Wendt, R. C., and Burwell, R. E. 1985. J. Soil Water Conserv. 40, 450-454. Wendt, R. C., and Corey, R. B. 1980. J. Environ. Qual. 9, 130-136. White, E. M. 1973. J. Environ. Qual. 2, 104-107. White, E. M., and Williamson, E. J. 1973. J. Environ. Qual. 2, 452455. White, E. M., Williamson, E. J., and Kingsley, Q. 1977. J. Environ. Qual. 6 , 251-254. White, R. E., and Beckett, P. H. T. 1964. Plant Soil 20, 1-16. Wildung, R. E., and Schmidt, R. L. 1973. Office of Res. and Monitoring, USEPA, Washington,

Wildung, R. E., Schmidt, R. L., and Gahler, A. R. 1974. J. Environ. Qual. 3, 133-138. Williams, C. H., and David, 0. J. 1973. Aust. J . Soil Res. 11, 43-56. Williams, C. H., and David, 0. J. 1976. Soil Sci. 121, 86-93. Williams, J. D., Shear, H., and Thomas, R. L. 1980. Limnol. Oceangr. 25, 1-11.

Buffalo, New York.

D. C. EPA-R3-73-04, pp. 185.