Phosphorus in Runoff and Streams

PHOSPHORUS IN RUNOFF AND STREAMS J. C. Ryden,' J. K. Syers,' and R. F. Harris Department of Soil Science, University of Wisconsin, Madison, Wisconsin

I. Introduction 11. Terminology

..................................................... . ...................... .............................

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B. Forms of 111. Factors Affecting

in Runoff and Streams

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B. Chemical Factors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . IV. Phosphorus Loads in Runoff and Streams . A. Influence of Point Sources on Phosphorus in Streams . . . B. Runoff from Forest Watersheds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C. Runoff from Agricultural Watersheds D. Runoff from Land Associated with Ani E. Urban Runoff . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . V. Impact of Phosphorus Carried in Streams on Standing Waters . . . . . . VI. Prcsent Status and Outlook . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

I.

1 2 2 4 4 4

7 20. 21 22 25 32 33 37 38 41

Introduction

Increasing evidence suggests that phosphorus ( P ) in surface waters is a primary factor controlling the eutrophication of water supplies (Ohle, 1953; Mackenthun, 1965; Stewart and Rohlich, 1967; Vollenweider, 1968; Lee, 1970). Assessment of the relative contribution of the different sources of P to surface waters (Fig. 1 ) is of critical importance for implementation of control measures to prevent or reverse P-induced eutrophication. Although the importance of runoff and streams as major sources of P to standing waters is well recognized, little attempt has been made to differentiate between and quantify the P forms in runoff and streams which are of potential importance with respect to their impact on the biological productivity of standing waters. Furthermore, little emphasis has been placed on the reactions that may occur between dissolved inorganic P and Present address: Department of Soil Science, Massey University, Palmerston

North, New Zealand.

1

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J. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

the solid phases with which it is in contact in runoff and streams, as pointed out by Taylor ( 1967) and Biggar and Corey (1969). Critical concentration limits have been suggested for P in surface waters which, if exceeded, will lead to excessive biological productivity (Sawyer, 1947; Mackenthun, 1968). In this review, however, rather than emphasizing critical concentrations, P in runoff and streams will be discussed mainly from the standpoint that any P load constitutes a potential increase in the P fertility of surface waters. II.

Terminology

A. HYDROLOGY AND PHOSPHORUS SOURCES This review will use essentially the definitions proposed by Langbein and Iseri ( 1960). Watershed (drainage basin; catchment area). A part of the surface of the earth that is occupied by a drainage system, which consists of a surface stream, or a body of standing (impounded) surface water, together with all tributary surface streams and bodies of standing surface water. Stream. A general term for a body of flowing water. In hydrology the term is usually applied to the water flowing in a natural channel. Stream flow. The discharge (of water) that occurs in a natural channel. Runoff.That part of precipitation that falls on land and ultimately appears in surface streams and lakes. Runoff may be classified further according to its source. Surface runoff (overland flow). That part of rainwater or snowmelt which flows over the land surface to stream channels. Surface runoff may also enter standing waters directly or be consolidated into artificial channels, e.g., storm sewers in urban areas (urban runoff), before entering a stream or body of standing water. Subsurface runoff (storm seepage). That part of precipitation which infiltrates the surface soil and moves toward streams as ephemeral, shallow, perched groundwater above the main groundwater level. In many agricultural areas subsurface runoff may be intercepted by artificial drainage systems, e.g., tile drains, accelerating its movement to streams. Groundwater run08 (base runoff). That part of precipitation that has passed into the ground, has become ground water, and is subsequently discharged into a stream channel or lake as spring or seepage water. In addition to runoff, the other potential contributors to streams and standing waters are precipitation incident on the water surface and industrial and sewage effluents (Fig. 1 ) .

PHOSPHORUS IN RUNOFF AND STREAMS

3

McCarty (1967) and Vollenweider ( 1968) have made a useful division of sources of P to surface waters based on the ease of quantification. Point sources enter at discrete and identifiable locations and are therefore amenable to direct quantification and measurement of their impact on the receiving water. Major point sources include effluents from indus-

FIG. 1. Schematic representation of the relationships between phosphorus sources and runoff, streams, and standing waters.

trial and sewage-treatment plants (Fig. 1) . Diffuse .wurces may be defined as those which at present can be only partially estimated on a quantitative basis and which are probably amenable only to attenuation rather than to elimination. Diffuse sources require the most investigative attention. Vollenweider ( 1968) further divided diffuse sources into: 1. Natural sources such as eolian loading, and eroded material from virgin lands, mountains and forests. 2. Artificial or semiartificial sources which are directly related to human activities, such as fertilizers, eroded soil materials from agricultural and urban areas, and wastes from intensive animal rearing operations. The loads of P imparted to runoff and streams from natural diffuse sources provide a datum line against which the magnitude of P loads from artificial sources may be compared.

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J. C. RYDEN, J. K.

SYERS, AND

R. F. HARRIS

B. FORMS OF PHOSPHORUS

In natural systems, P occurs as the orthophosphate anion (Pod3-)which may exist in a purely inorganic form (H2P0,- and HP0,2-) or be incorporated into an organic species (organic P ) . Under certain circumstances inorganic orthophosphate may exist as a poly- or condensed phosphate. A secondary distinction is made between particulate and dissolved forms of P, the split conventionally being made at 0.45 pm. Other terminology used is as follows: Total P . All forms of P in a runoff or stream sample (dissolved and particulates in suspension) as measured by an acid-oxidation treatment (e.g., acid ammonium persulfate). Dissolved inorganic P . P in the filtrate after 0.45 pm separation determined by an analytical procedure for inorganic orthophosphate. Organic P . P that may be determined within the dissolved and particulate fractions by the difference between total P and inorganic P. Ill.

Factors Affecting the Dynamics of Phosphorus in Runoff and Streams

Before evaluating the magnitude of various P sources in terms of the loads of P in runoff and streams, and the extent to which previous studies of P loadings enable an adequate definition of P sources, it is important to understand the physical and chemical factors affecting the dynamics of P in runoff and streams. These factors determine not only the movement of P into runoff and streams, but also its distribution between the aqueous and particulate phases. A.

PHYSICALFACTORS

All terrestrially derived diffuse sources of P are associated with the movement of water in contact with a solid phase. The solid phase may be stationary with respect to water flow, or may move in the flow at some speed equal to or less than the flow. Precipitation disposed of as subsurface or groundwater runoff is primarily in contact with a stationary solid phase, namely the soil profile and, in the case of groundwater runoff, possibly the bedrock. Consequently, the amounts and concentrations of P carried in subsurface and groundwater runoff will be influenced by the time of contact with any component in the soil profile capable of interacting with dissolved P in the percolating water and by the concentration of dissolved P that the soil components maintain in the soil solution. Time of contact between the percolating solution and any soil component will in turn depend on the rates of infiltration and percolation into and through the soil.

PHOSPHORUS IN RUNOFF AND STREAMS

5

Some of the theories developed to describe water movcment in soils can be applied to evaluate the potential loss of P from various soil types as a result of subsurface runoff. Gardner (1965) developed equations to describe the movement of nitrate in the soil profile due to leaching. The chemical interactions that occur between dissolved inorganic P and soil components (discussed later), when water percolates through the soil, must also be taken into consideration. Inclusion of a term in the equations developed by Gardner (1965) to describe the relationship between P in particulate and aqueous phases is therefore necessary. This could take the form of a linear adsorption isotherm relevant to the concentrations of dissolved inorganic P maintained in the solution of a particular soil. Biggar and Corey (1969) have also reviewed the literature on infiltration and percolation of water in agricultural soils as it pertains to nutrient movement. The movement of solid phase material in contact with natural waters occurs during surface runoff and in streams. The amounts of solid material capable of entering surface runoff will depend on the intensity of rainfall, physical and chemical attachment between various solid components, and the amounts and energy of runoff waters (Guy, 1970). It is the energy of surface runoff or stream water, however, that governs the amounts of a specific size fraction of particulate materials which will remain in suspension during water flow. The primary source of particulate material to surface runoff and streams is eroding soil (Guy and Ferguson, 1970), although in urban areas with little ongoing development, particulates may be dominated by specifically urban detrital material (e.g., street litter and dust) and organics derived from urban vegetation. The total on-site losses of soil due to sheet and rill erosion are not necessarily delivered to streams. The amount of sediment that travels from a point of erosion to another point in the watershed is termed the sediment yield (Johnson and Moldenhauer, 1970). Consequently the Universal Soil Loss Equation used to predict field soil losses on an average annual basis (Wischmeier and Smith, 1965) must be corrected when used to predict sediment loads in streams because deposition of particulates may occur on the land surface as a result of slope variations before surface runoff reaches a stream. It is for this reason that estimates of soil loss in surface runoff from sites within a particular watershed cannot be translated into total P losses through a knowledge of the total P content of the soil, if the P loss is to be related to P enrichment of surface waters. An associated complication arises from the fact that soil P is primarily associated with the solid phase. As soil erosion is a selective process with respect to particle size, selectivity has been observed for P loss in surface

6

J. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

runoff. The extent of the selectivity depends on the particle sizes with which most of the soil P is associated. This observation has led to the concept of enrichment ratios (ER) , which for P are calculated as the ratio of the concentration of P in the particulate phase of surface runoff to the concentration of P in the source of the particulate phase. This effect was first considered by Rogers (1941), who observed ER values of 1.3 for total P and 3.3 for “0.002 N H,SO, extractable” P for a silt loam situated on a 20-25% slope. Other values range from 1.5 to 3.1 for total P (Knoblauch et al., 1942; Neal, 1944; Stoltenberg and White, 1953), 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. The selective nature of surface runoff with respect to P is due to selective removal of fine soil particulates as a result of the energy limitations of runoff and the fact that a large percentage of total soil P is frequently associated with clay-sized material (Scarseth and Chandler, 1938; Williams and Saunders, 1956; Syers et al., 1969). Greater selectivity of fines and consequently particulate P will occur as the energy of surface runoff decreases. Stoltenberg and White (1953) observed that as precipitation disposed of through surface runoff decreased from 70 mm to 0.25 mm per hour, the clay content of eroded material from a soil with a clay content of 16-18% increased from 25% to 60%. This has obvious implications in relation to the nature of the sediment load carried by a stream and the interactions of P between the solid and aqueous phases, particularly during periods of surface runoff. It should be pointed out, however, that although the P content of the sediment load may increase as surface runoff diminishes, as may be predicted from the work of Stoltenberg and White (1953), the total P load may not change, or may even decrease, owing to lower sediment loads. The particulate material carried in streams may be divided into bed load and wash or suspended load. The bed load, which may also have a contribution from existing stream sediment, is that which moves along or close to the stream bed, whereas the wash load is maintained in the flow by turbulence (Johnson and Moldenhauer, 1970). By inference from the selectivity of surface runoff for fine soil particulates, the wash load will be high during surface runoff events. Furthermore, Johnson and Moldenhauer (1970) suggested that the wash load travels at about the same velocity as the water with which it is in contact. Consequently, P associated with the clay- and silt-sized particulates constituting the wash load will move between any two points in the stream profile at the same speed as the ambient dissolved forms of P. Increased turbulence in streams during high flow, or arising from an increasing gradient, will tend to maintain in suspension particle sizes more

PHOSPHORUS IN RUNOFF AND STREAMS

7

characteristic of the bed load, and may even resuspend existing stream bed sediment. In a study of total P loads in the Pigeon River, North Carolina, Keup (1968) noted that an increase in gradient from 2.81 to 4.35 m/km, over which no tributaries entered the main stream, resulted in a 90.8 kg/day increase in the total P load carried. It appears that in the majority of cases a large proportion of particulate P in streams arises from soil erosion. Phosphorus may be stored in stream bed sediments, but unless the stream is actively aggrading, the amount of P stored will be less than the inflow (Keup, 1968). That which is stored is liable to resuspension and transport owing to turbulence during periods of high flow.

B. CHEMICAL FACTORS 1 . Nature of Soil P Soil P may be divided into two broad categories: inorganic P, namely, that associated with soil mineral particles; and organic P, which forms an integral part of the soil organic matter fraction. a. Inorganic P . O n the basis of solubility product criteria, it has been postulated that discrete phase crystalline Fe and A1 phosphates exist in noncalcareous soils (Kittrick and Jackson, 1956; Hemwall, 1957; Chakravart and Talibudeen, 1962). The general occurrence of discrete Fe and A1 phosphates seems doubtful on the basis of the ion product data presented by Bache (1964) and the experimental observations of Hsu (1964). It is now generally accepted that secondary inorganic P in many soils exists primarily in association with oxides and hydrous oxides of Fe and Al, as surface-bound forms or within the matrices of such components. However, that discrete Fe and A1 phosphates are formed as temporary phases in the vicinity of phosphate fertilizer particles due to conditions of localized high acidity and P concentration is well established (Lindsay and Stephenson, 1959; Huffman, 1969). Such compounds will not be stable as the dissolved inorganic P concentration in the soil solution or aqueous portion of other soil-water ecosystems decreases. The calcium phosphate mineral, apatite (Shipp and Matelski, 1960) and calcic fertilizer-soil reaction products (Huffman, 1969) have been identified in soils. The amounts of apatite are appreciable only in weakly weathered soils (Williams et al., 1969), as predicted by the weathering indices of Jackson ( 1969). Calcic fertilizer-sail reaction products may be present in neutral and calcareous surface soil horizons, and their importance in maintaining high concentrations of dissolved inorganic P in soil-water ecosystems should not be overlooked.

8

J. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

Consequently three basic forms of inorganic P may exist in unfertilized soils (Syers and Walker, 1969; Williams and Walker, 1969): apatite, which is a discrete phase P compound; P sorbed on the surfaces of Fe, Al, and Ca soil components (nonoccluded); and P present within the matrices of Fe and A1 components (occluded). In fertilized soils, a variety of P fertilizer-soil reaction products may exist as transient phases. As the solubility product of pure apatite in water is low (0.03 pg per milliliter at pH 7, Stumm, 1964) and the P held within the matrices of Fe and A1 components is virtually chemically immobile, except under reducing conditions in the case of Fe, major emphasis should be directed toward the reactions involving P in solution and that sorbed on the surfaces of Fe, Al, and Ca components as well as the release of P due to dissolution of fertilizer-soil reaction products. b. Organic P. Elucidation of the composition of soil organic P is restricted by lack of extractants capable of removing organic P from soils in a relatively unaltered form and by the inadequacy of current methods for mildly degrading extracted organic P-organic matter complexes. Existing data indicate that most of the organic P in soils is associated, in an ill-defined manner, with the humic and fulvic acid complex of soil organic matter (Anderson, 1967). Of the specific forms of organic P that have been identified in soils, inositol phosphates are present in largest relative amounts, comprising up to 60% of the total organic P (Anderson, 1967; Cosgrove, 1967; McKercher, 1969). Other specific organic P compounds are present in soil in much lower quantities: nucleic acids account for 5-lo%, and other phosphate esters, such as phospholipids, sugar phosphates, and phosphoproteins, for less than 1-2% (McKercher, 1969). 2. Sorption of Dissolved P by Soils Whenever water containing a particular concentration of dissolved P comes into contact with soil material, there is a possibility for sorption, desorption, or dissolution reactions to take place. The types of reactions are the same regardless of whether they occur under conditions existing in the soil profile, surface runoff, or streams. Although in some cases biological assimilation may initially affect the distribution of P between dissolved and particulate phases of soil-water systems, the distribution of P between these phases will be determined by the nature of the inorganic particulates and the concentrations of dissolved P in solution (Keup, 1968; McKee et al., 1970; Ryden et al., 1972b). a. Inorganic P. It has been demonstrated that the uptake or sorption of P from solution by soils is significantly related to the presence of shortrange order (amorphous) oxides and hydrous oxides of Fe and A1 (Williams et al., 1958; Gorbunov et al., 1961; Bromfield, 1965; Hsu, 1964; Saunders, 1965; Syers et al., 1971). Furthermore, “pure” oxides and hy-

PHOSPHORUS IN RUNOFF AND STREAMS

9

drous oxides of Fe and Al, and short-range order aluminosilicates have also been shown to be particularly effective in the sorption of inorganic P from solution (Gastuche et al., 1963; Muljadi et al., 1966; Hingston et al., 1969). The sorption of inorganic P by Fe and A1 oxides and hydrous oxides is known to be rapid, as is the sorption of P by soils. Furthermore, short-range order Fe and A1 oxides and hydrous oxides are ubiquitous in soils (Hsu, 1964), their relative amounts depending on parent material, climatic and drainage conditions, and occur mainly as coatings on other soil components. Shen and Rich (1962) and Jackson (1963) have noted the occurrence of A1 hydroxypolymers and Dion (1944), and Roth et al. ( 1969) have reported the presence of F e oxide and hydrous oxide coatings on clay mineral surfaces. Such coatings, in conjunction with the greater surface area of the clay fraction compared to that of the other particle-size fractions in a soil, explain the observation of Scarseth and Chandler (1938) that up to 50% of the total P in soils may be associated with the the clay fraction, as well as the enrichment ratio effect for P as a result of soil erosion. Attempts have been made to correlate P sorption with the clay content of soils (Williams et al., 1958). Correlations between P sorption and clay content after removal of Fe and A1 oxides and hydrous oxides often have been poor. Better correlations may be expected if P sorption is related to the content of water-dispersed clay. The sorption of P by water-dispersed clay and silt of soils has obvious implications to reactions occurring between dissolved and particulate P in surface runoff and streams. Sorption of inorganic P by CaC03 has also been demonstrated (Cole et al., 1953). The nature of the surfaces of calcite in calcareous soils may be very different from those of pure calcite (Buehrer and Williams, 1936; Lahav and Bolt, 1963; Syers et al., 1972). The sorption of dissolved inorganic P by soils may be described by sorption isotherms similar to that shown in Fig. 2. Numerous workers have also shown that sorption may be described by some of the adsorption isotherms developed to describe gas adsorption by solids (Russell and Prescott, 1916; Olsen and Watanabe, 1957; Rennie and McKercher, 1959; Syers et al., 1973). Similar observations have been made for the sorption of inorganic P by soil components such as kaolinite and short-range order Fe and A1 oxides and hydrous oxides (Gastuche et al., 1963; Muljadi et al., 1966; Kafkafi et al., 1967). Although these studies have been useful in describing relationships between various soils and soil components with respect to their P sorption capacities, they have provided little information regarding P sorption behavior from solutions containing the low dissolved inorganic P concentrations characteristic of most soil-water ecosystems, largely because of the high levels of added P used (Ryden et al., 1972b). Furthermore, Syers et al. (1973) obtained two linear Langmuir relation-

10

J. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

sorbed ( f )

APon sol I

released (3

FIG.2. Typical isotherm for the sorption of added inorganic phosphorus by a soil. E = equilibrium P concentration. (From White and Beckett, 1964.)

ships which intersected at equilibrium P concentrations varying from 1.5 to 3.2 pg P/ml, for three contrasting soils-an observation that probably invalidates interpretations of P sorption made from many previous studies where high levels of added P were used. The study of White and Beckett (1964), conducted at initial dissolved inorganic P concentrations, comparable to those existing in soil-water ecosystems, provides a useful basis for understanding the interactions between aqueous and particulate phases of P in runoff and streams. Figure 2 illustrates the principle of the approach used. White and Beckett (1964) defined the intersection of the P sorption isotherm and the abscissa, the “equilibrium phosphate potential” ( 5 p C a pH,PO,) , abbreviated to “equilibrium P concentration” by Taylor and Kunishi ( 1971) . The intersection is equivalent to the inorganic P concentration in the ambient aqueous phase when there is no net sorption or release of P, i.e., AP = 0. This is a point of reference which provides a predictive estimation of sorption or release of P should the P concentration in solution change. Furthermore, the average slope of the sorption curve over a given final P concentration range provides information on the ability of the soil to maintain the P concentration at the equilibrium P concentration. The steeper the slope, the closer will the final P concentration be to the equilibrium P concentration. The slope of the curve, although not related to total P sorbed, is related to the extent to which that soil may sorb P over the concentration range considered. The potential of this approach in predicting the chemical mobility of P in soil-water systems is clearly evident and has been used with regard to streams by Taylor and Kunishi (1971) and Ryden et al. (1972a,b) for rural and urban soils, respectively. The desorption of sorbed P from soils is not as simple as may be inferred from the sorption-release relationships obtained by White and

+

PHOSPHORUS I N RUNOFF AND STREAMS

11

Beckett (1964). In fact very few studies have been reported regarding the desorption of sorbed P, and those reported by Syers et al. (1970) and Ryden et al. (1972a), involved desorption following sorption of P from solutions containing P concentrations in excess of those commonly found in soil-water ecosystems. In studies involving the sorption of P by kaolinite from solutions containing realistic inorganic P concentrations, Kafkafi et al. (1967) observed that initially all the sorbed P was isotopically exchangeable. During a subsequent washing or desorption step, however, a portion of the sorbed P became nonexchangeable, or “fixed,” this portion being dependent upon the amount of P sorbed, the number of washings, and the nature of the previous P sorption cycle. Sorption of P was represented by either onestep sorption from a range of solutions of different initial P concentration or by successive additions of small amounts of dissolved inorganic P. Both these types of P sorption, as well as an effect analogous to washing, could occur in soil-water ecosystems. 6. Organic P . Although the mechanisms involved in the retention of organic P by soils have not been established fully, there is evidence that inositol hexaphosphate, and possibly other organic P compounds, are retained by a precipitation rather than a sorption reaction. Nevertheless, removal of dissolved organic P from solution appears to be a rapid process. Pinck et al. (1941 ) reported that many commonly occurring water-soluble organic phosphates, e.g., salts of glycerophosphate, hexose diphosphate, and nucleic acids, become nonextractable with water at almost the same rate and as completely as dissolved inorganic P. The retention of water-soluble organic P by sorption reactions may occur by at least two basically different mechanisms (Sommers et al., 1972). Goring and Bartholomew (1950) observed that removal of “free iron oxides” considerably reduced the amount of fructose 1,6-diphosphate sorbed by subsoil material, suggesting that the sorption of organic P may occur through orthophosphate groups by a similar mechanism to that for inorganic P. It is also possible that organic P can be retained by interaction of the organic moiety of the phosphate ester with inorganic soil components. For example, nucleic acids and nucleotides are protonated at pH 5 (Jordan, 1955) and could consequently be retained on clay surfaces by displacement of exchangeable cations. Furthermore, physical adsorption, also through the organic portion of the molecule, is possible, particularly if the molecular weight of the compound is high, as suggested by Greenland (1965). In such cases retention is weak and is accomplished by van der Waals and ion-dipole forces. Greaves and Wilson (1969) have implicated physical adsorption in the retention of nucleic acids by montmorillonite. It is also possible that retention occurs indirectly through other

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J. C . RYDEN, J. K. SYERS, AND R. F. HARRIS

soil organic compounds such as fulvic and humic acids after interaction of organic phosphates with these species (Martin, 1964). The desorption of sorbed organic P has not been extensively studied. The hypothesis that inorganic P added to soils displaces sorbed organic P to solution (Latterell et al., 1971) was not supported by the data presented by Wier and Black (1968). Although organic P may be leached from soils, it appears that a large proportion of that removed may not be in a dissolved form. After incubating sucrose with ammonium nitrate in the upper portion of a calcareous soil, Hanapel et al. (1964) found that most of the organic P removed by leaching was present in a particulate rather than a dissolved form. 3. Chemical Aspects of P in Subsurface and Groundwater Runoig Losses of P in subsurface and groundwater runoff have been considered minimal in the past, but, as will be discussed later, such losses can amount to a significant proportion of losses from agricultural land, and possibly a major proportion from forest lands. The supposition that P losses in subsurface and groundwater runoff are low probably stems from the concept of P immobility based on the P sorption properties of soils using added inorganic P concentrations far in excess of those normally present in the soil solution. It is of interest to note that many of the reported mean concentrations of dissolved inorganic P in subsurface runoff are within the range of values expected to be maintained in the soil solution. Pierre and Parker (1927) reported values ranging from 0.020 to 0.350 pg P/ml, with an average of 0.090 pg/ml, for several surface soils from the southern and midwestern states of the United States. These workers also noted that dissolved inorganic P concentrations could be maintained at a fairly constant level. Barber et al. (1963) reported similar values for the upper 15 cm of 87 soils from the midwestern United States, with an average of 0.180 pg of P per milliliter; the frequency distribution of the values obtained, however, suggested a mode of between 0.040 and 0.060 pg of P per milliliter. As water percolates through the soil profile, there tends to be a “chemical sieving” of dissolved inorganic P (Black, 1970). This arises as a result of the sorption of inorganic P by soil components. The low concentrations of P found in groundwater runoff, which has experienced the maximum effects of deep percolation with concomitant increase of contact with P-deficient particulates of the subsoil, are undoubtedly a direct result of the chemical sieving effect. The principle of this effect is illustrated by other data presented by Barber et al. (1963). For the same 87 soils mentioned previously, the average dissolved inorganic P concentration at a depth of 46-61 cm was 0.089 pg/ml, less than half that for the upper

PHOSPHORUS IN RUNOFF AND STREAMS

13

0-15 cm. Another illustration is observed in results presented by Ryden et al. (1972a) for the P sorption properties of successive soil horizons of a Miami silt loam profile. The concentrations of dissolved inorganic P maintained in solution after shaking with solutions of different initial added inorganic P concentrations at a solution: soil ratio of 40: 1 are given in Table I. TABLE I

Dissolved Inorganic Phosphorus (P) Concentrations Maintained by Soil IIorizoiis of Miami Silt Loam after Equilibration with Solutions of Different Initial Added Inorganic P Concentrationsn

Horizon

Depth (cm)

Initial P conc. (ccg/ml)

Final P coiic. (/*g/mU

A1

0-15 15-38 56-66

0.0 0.471 0.030

0.471 0.030 0.007

B1 3C1

~

Data extrapolated from Ryden et

(11.

(197‘2a).

The concentration of dissolved inorganic P in subsurface and groundwater runoff will depend on the nature and amounts of P-retaining components in the profile, the surface area exposed to percolating waters, and the ease of percolation which affects the contact time of dissolved inorganic P with the retaining components. In studies of P leaching through columns of organic soils in the laboratory, Larsen et al. (1958) observed that P retention, measured by srP autoradiographs, was closely correlated with the total hydrous Fe and A1 oxide (“sesquioxide”) content. Similarly, losses of P due to leaching through a deep siliceous sandy soil were demonstrated in W. Australia by Ozanne (1963). When 225 kg/ha of 32P-labeled superphosphate was broadcast during winter on a fallow sandy soil, over 50% of the P had penetrated to more than 1 m below the surface within 38 days, during which 230 mm rain had fallen. Ozanne (1963) also demonstrated that the potentially large losses of P to subsurface and groundwater runoff from sandy soils compared to that from loamy soils were due to quantitative rather than qua1itativ.e differences in P-retaining components. Although major emphasis has been placed on P losses in surface runoff, it appears that losses of P to subsurface and groundwater runoff, although of little significance from an agricultural standpoint, may under certain conditions constitute a significant loss of P from agricultural watersheds in terms of the P enrichment of surface waters, as will be discussed

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J. C. RYDEN, J . K . SYERS, AND R. F. HARRIS

later. Losses of P to subsurface and groundwater runoff are even more difficult to evaluate than those in surface runoff and demand further investigative attention.

4 . Chemical Aspects of P in Streams As discussed previously, surface runoff from agricultural land constitutes a heterogeneous and relatively short-lived system. Any attempt to consider the distribution and chemical mobility of P between solid and aqueous phases before entry into the receiving stream would be pointless as a new and more homogeneous system is rapidly established. Surface runoff in urban areas is somewhat different because in most cases it is channelized shortly after origin by alteration of surface drainage patterns; under such circumstances it is analogous to a stream in an artificial channel. Consequently, the chemical mobility of P will be discussed from the standpoint of the stream environment. The potential of suspended particulates derived from eroding soil to modify the dissolved inorganic P concentration of streams has been suggested by Taylor ( 1967) and Biggar and Corey (1969). Wang and Brabec (1969) also implied that inorganic P was sorbed by suspended particulate material from observations of dissolved inorganic P concentrations in the Illinois River at Peoria Lake. An evaluation of the possible effects of eroded soil materials on the dissolved inorganic P concentrations of streams may be obtained from P sorption studies (Taylor and Kunishi, 1971; Ryden et al., 1972a,b). It is essential, however, that conditions realistic of those existing in streams are used if meaningful results are to be obtained (Ryden et al., 1972a). Widely differing interpretations can be made as solution: soil ratios and initial dissolved inorganic P concentrations are changed from those conventionally used in P sorption studies to those realistic in terms of the stream environment (Fig. 3 a-c). The data in Fig. 3a suggest that inorganic P released from the A1 horizon, which contained a P fertilizer-soil reaction product, would be largely resorbed by the noncalcareous B1 horizon and to some extent by the calcareous 3C1 horizon, should the horizons erode together. Sorption studies employing low initial added inorganic P concentrations and a wide (400: 1 ) so1ution:soil ratio (Fig. 3c) indicate that the B1 horizon has a much lower ability to remove dissolved inorganic P from solution than expected, this being equal to or only slightly greater than that of the 3C1 horizon. In fact for mixtures of varying ratios of A1 and B l , and A1 and 3C1 horizons, it was found (Ryden et al., 1972b) that the latter mixtures were able to maintain lower dissolved inorganic P concentrations than the former. The conditions used by Ryden et al. (1972a,b) to predict the potential of eroding soils to modify the dissolved

PHOSPHORUS IN RUNOFF AND STREAMS

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.30

10

10

0

-

10

Final dissolved inorganic P Concentration Wgll)

FIG. 3. Sorption of added inorganic phosphorus by horizons of a Miami silt loam profile from solutions of varying initial dissolved inorganic P concentrations and at varying so1ution:soil ratios. ( a ) High added P (0-6 pg/ml) and narrow so1ution:soil ratio (50: 1 ) . ( b ) Low added P (0-0.2 pg/ml) and narrow solution:soil ratio ( 4 0 : l ) . (c) Low added P (0-0.2 pg/ml) and wide so1ution:soil ratio ( 4 0 0 : l ) . [From Ryden et al. (1972a), reproduced with permission of the American Society of Agronomy.]

inorganic P concentrations of streams, gave results comparable to those obtained in simulated stream systems using a solution: soil ratio of 1000:1 This is equivalent to a sediment concentration of 1000 mg/liter, which lies well within the range of values cited by Guy and Ferguson (1970) and Johnson and Moldenhauer ( 1970). The P sorption studies reported by Taylor and Kunishi (197 1) and Ryden et al. (1972a,b) involved closed systems, i.e., soil in contact with the same aqueous phase. This may be justified on the grounds that the wash load of a stream travels at the same velocity as the water in which it is suspended (Johnson and Moldenhauer, 1970), as discussed previously.

16

J. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

Sorption studies may be used to provide reasonable estimates of dissolved inorganic P concentrations in streams, under various flow conditions, draining rural watersheds. Taylor and Kunishi (1971 ) observed that dissolved inorganic P concentrations during base flow of a stream draining a small agricultural watershed in Pennsylvania, were in the range of 0.040 to 0.060 pg P/ml, values which were close to those predicted from P sorption studies using stream bank sediment and subsoil material. During periods of surface runoff, predicted dissolved inorganic P concentrations would be in excess of 0.200 pg of P per milliliter for the surface soil used by Taylor and Kunishi (1971) and 0.100 pg of P per milliliter for that used by Ryden et al. (1972a) due to release of P from eroded surface soil; however, predictions from the work of Taylor and Kunishi (1971) are based on the use of a narrow (10: 1 ) so1ution:soil ratio. The ability of eroding stream bank material or resuspended stream bed sediment to resorb inorganic P released to solution should not be ignored (Taylor and Kunishi, 1971 ) . In a more recent study of the same watershed in Pennsylvania, Kunishi et al. (1972) observed that during a heavy summer rainstorm only 31% of the total “available” P (total dissolved plus resin-extractable P on the suspended sediment) was in the resin-extractable form in a stream draining an agricultural subwatershed. At the outflow of the main watershed, however, over 50% of the total “available” P was in the resin-extractable form. Kunishi et al. (1972) suggested that for this watershed, as suspended material moves downstream and mixes with material from other parts of the watershed as well as that eroded from the stream banks, dissolved P is actively sorbed. During a second less intense storm, however, when stream bank erosion was less severe, the proportion of total “available” P associated with the sediment was virtually the same at both monitoring stations. A similar hypothesis might also explain the observation of White ( 1 972) at Taita, New Zealand, that the concentration of dissolved inorganic P at the outflow of small watershcds during base flow was lower than that recorded for groundwater seepage giving rise to the stream flow. It is important to distinguish between the quantities of various types of soil materials expected to enter streams in urban as opposed to agricultural surface runoff. In agricultural areas, surface runoff will carry primarily surface soil material to receiving streams. Surface soils may contain P fertilizer-soil reaction products capable of producing significant increases in dissolved inorganic P concentrations, due to their dissolution (Ryden et al., 1972a). In urban areas, however, land under development, which is prone to severe erosion, is frequently graded, exposing some or all horizons of the area profile to potential erosion. Dissolved inorganic P concentrations of receiving streams in urban areas may be sufficiently high that

PHOSPHORUS IN RUNOFF AND STREAMS

17

the addition of eroded soil material may cause a reduction in the dissolved inorganic P concentration. An approach similar to that used by Taylor and Kunishi (1971) and Ryden et al. (1972a,b) could be used to identify other diffuse sources of potential P enrichment within a watershed. The approach would be particularly useful for estimating the potential of various forms of urban detrital material to influence the dissolved inorganic P concentrations of surf ace runoff. One diffuse source of considerable importance is the leachate from leaves, particularly during the autumn. An appreciable percentage of the total P in leaf tissue may be in a water-soluble form. Ash leaves may contain 62% of total P as water-soluble inorganic P (Nykvist, 1959). Cowen and Lee (1972) observed that 44 and 120 pg soluble inorganic P per gram air-dry weight of fallen oak and poplar leaves, respectively, could be leached by 1 liter of distilled water percolating at a rate of 8.4 ml per minute. Greater amounts of P were released from oak leaves during consecutive leaching cycles and after fragmentation of whole leaves. Similar experiments were conducted by Timmons et al. (1970) using agricultural crop residues. These were leached in a fresh condition and after drying, and freezing and thawing cycles. The data suggest that the leaching of crop residues is most likely to contribute to the dissolved inorganic P concentration of streams during spring thaw in certain areas when, after numerous freezing and thawing cycles, the residues will be carried over frozen ground in surface runoff. When greater infiltration can occur, a portion of the leached P may be retained in the soil due to sorption.

5 . P Chemistry of Stream-Bed Sediment Little is known of the chemistry of stream-bed sediment although it is conceivable that it is similar to that of the subsoil of the surrounding area (Taylor and Kunishi, 1971 ) . Consequently, P sorption studies using subsoil material may provide some information on the role of stream-bed sediment in regulating the dissolved inorganic P concentration due to its suspension during turbulence. This would be particularly true in watersheds with little contribution to stream-bed sediment as a result of surface runoff. In watersheds where surface runoff is a regular occurrence, however, stream-bed sediment is expected to have a significant contribution from surface horizon soil material, and the latter could contribute to base flow concentrations of inorganic P. Care should be taken, however, in the extension of the P sorption properties of field soils to stream-bed sediment. Hsu (1964) observed that the amount of inorganic P sorbed by soil after storage for 1 year in a continuously wet condition, increased from 69 to 99 pg of P per gram of soil.

18

J. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

The increased sorption was attributed to release of Fe to solution from crystalline phases due to the development of localized reducing conditions during storage, and reprecipitation of “ferric hydroxide” on contact with more aerobic conditions. The redox status of stream-bed sediments does not appear to have been studied, but it is reasonable to suggest that reduction occurs at depth in the sediment with the possibility of crystalline ferric components being transformed to short-range order ferrous forms. The importance of short-range order oxides and hydrous oxides of Fe in the sorption of inorganic P has already been discussed. The possible transformation of Fe from crystalline to short-range order forms represents the first stage of the more aggressive transformations which occur in lake sediments under anaerobic conditions (Shukla et al., 1971). The observation of Kafkafi et al. (1967) that the washing of kaolinite, on which P had been sorbed, produces a “pool” of nonexchangeable P is also of direct relevance to the P chemistry of stream-bed sediments, assuming a similar effect occurs. Stream-bed sediment with associated sorbed P could undergo a series of steps equivalent to sequential washing due to resuspension and settling as a result of minor turbulence. The observations of Kafkafi et al. (1967) suggest that sorbed P could become progressively less exchangeable and may constitute an essentially permanent removal of dissolved inorganic P from streams. When stream-bed sediment contains eroded fertilized soil materials, however, a different situation may prevail. Ryden et al. (1972a) showed that release of P from a surface soil horizon by repeated washing with P-free 0.1 M NaCl initially followed first-order kinetics, suggesting that release was due to the dissolution of solid phase P, probably a fertilizer-soil reaction product. 6 . Forms of P in Runoffand Streams

In many studies concerned with various aspects of P in runoff and streams there has been a tendency to measure total P. The measurement of total P discharged by streams does not provide any indication of the amounts of P available for aquatic plant growth. Consequently, the forms of P measured in streams that enter a lake or reservoir are of direct importance in assessing the impact of runoff- and stream-derived P on a body of standing water. Dissolved inorganic P is one of the obvious choices because this form of P is directly available for biological utilization. Objections to the measurement of dissolved inorganic P, as it is conventionally determined, have been raised by Frink (1971 ) on the basis that distinction between dissolved and particulate forms is based on filtration through a 0.45 pm filter. Although it is possible that filtration does not strictly differentiate between dissolved and particulate P, it provides a more

PHOSPHORUS IN RUNOFF AND STREAMS

19

realistic measure in terms of the effects of runoff- and stream-derived P on the biological productivity of standing waters than the measurement of total P. Vollenweider ( 1968) has also pointed out the necessity to distinguish between total P and dissolved forms of P because it is possible that P exports from some watersheds occur mainly in biologically unavailable forms, such as apatite. This work showed that P exports from the Alpine portion of the Rhine Basin amount to 1.45 kg/ha per year. As this is mainly in the form of apatite, however, the contribution of biologically available P to Lake Constance is relatively small. In other regions it appears that a high proportion of particulate inorganic P in streams draining urban and rural watersheds may in fact be apatite. Eroding urban soils in the Lake Mendota watershed, Wisconsin, contain between 6 and 80% of the total inorganic P as apatite, with amounts exceeding 50% in the lower B and C horizons (J. K. Syers, J. C. Ryden, and J. G. Thresher, unpublished data). For the same soil materials, Sagher and Harris (1972) observed that algal cultures suffered P starvation when the sole P source in the growth medium was C horizon material, indicating the very low biological availability of the P present in apatite. Chemical fractionation schemes have been used to determine the forms of inorganic P in soils. These schemes evolved from the observations of Chang and Jackson (1957) that certain chemical reagents were able to solubilize inorganic P contained in various synthetic phosphates and phosphate minerals. Recent workers (Bromfield, 1967; Williams et al., 1967, 1971a,b; Syers et al., 1972) have questioned the validity of the separation of inorganic P into Al-, Fe-, and Ca-bound forms, as proposed by Chang and Jackson ( 1957). Providing that the problems inherent in inorganic P fractionation schemes are recognized, useful interpretations may be drawn from the data obtained. The form of particulate inorganic P which is expected to have the greatest potential impact on the biological productivity of standing waters is that which is nonoccluded. Part of the nonoccluded and even some of the occluded inorganic P associated with ferric components is released into solution when anaerobic conditions develop subsequent to sedimentation. Appropriate inorganic P fractionation schemes applied to suspended stream sediments may provide a more meaningful measure of the forms and amounts of particulate inorganic P carried in streams. As pointed out by Taylor et al. (1971), suspended sediment concentrations are frequently not high enough to provide adequate amounts of xaterial in a manageable volume of water. Evaluation of the forms of P in soil materials which are known to be transported to streams in surface runoff may overcome this problem to some extent. In the case of eroding soil materials, the inorganic P fractionation schemes should not

20

J. C. RYDEN, J. K. SYERS, A N D R. F. HARRIS

be applied to the whole soil, due to the ER effect resulting from erosion. Water-dispersed particle-size separates should be used. In spite of the possible errors involved in a dissolved-particulate P split based on filter pore size, it seems that in the majority of cases the most meaningful and useful measurements of P in runoff are dissolved forms, particularly dissolved inorganic P. Frequently dissolved forms of P account for a major percentage of total P (Sylvester, 1961; Sullivan and Hullinger, 1969), whereas dissolved inorganic P in many cases constitutes a major proportion of the total dissolved P. It should be noted that dramatic changes can occur in the concentration of dissolved inorganic P and other P fractions after sample collection, even after only a short period of time . some cases when samples are not analyzed im(Ryden et al., 1 9 7 2 ~ )In mediately after collection, the only valid P parameter that can be measured is total P. IV.

Phosphorus loads in Runoff a n d Streams

The P content of precipitation reflects the amount of P subject to washout from the atmosphere at the time of the precipitation event. The amounts of P carried in precipitation rarely exceed 1 kg/ha per year as total P (Miller, 1961; Weibel et al., 1966; Allen et al., 1968; Gore, 1968). Weibel et al. (1966) reported that the average concentration of total acidhydrolyzable P in precipitation falling on Cincinnati, Ohio, was 0.080 pg/ml, whereas Taylor et al. ( 197 1 ) reported an average concentration of 0.020 pg/ml for total dissolved P in precipitation collected at rural Coshocton, Ohio. Data for the P content of precipitation should be viewed with some skepticism unless adequate precautions have been taken to guard against contamination of the collection vessel (Gore, 1968; White, 1972). White (1972) found that although rainwater collected over an extended period indicated a mean dissolved inorganic P concentration of 0.020 pg/ml, a mean concentration of 0.003 pg/ml, based on specific showers, might be a more accurate estimate. It is difficult to evaluate the effect of P carried in precipitation on P loads in runoff and streams. Phosphorus contained in precipitation which becomes a part of any soil-water ecosystem may undergo considerable change in form, depending primarily on the chemical factors discussed previously, and will become an integral part of the P forms in runoff and streams. Surface runoff water is the carrier of not only the P initially present in precipitation but also any P which enters surface runoff water because

PHOSPHORUS IN RUNOFF AND STREAMS

21

of chemical interactions or the energy of the water itself. Several factors affect the amount and energy of surface runoff water at any particular location and, therefore, the amount of additional P entering and carried by it. These include nature of land use, extent of vegetative cover, slope, intensity of rainfall, and permeability of the land surface. The quantity of precipitation entering subsurface and groundwater runoff is inversely related to that disposed of in surface runoff and evapotranspiration. It is consequently affected by the factors listed above for surface runoff. The major portion of P in subsurface and groundwater runoff is expected to be in dissolved forms. If subsurface runoff is accelerated by artificial drainage systems, however, soil colloids, with associated P, may appear in the water as it enters streams. The P load carried by a stream under given flow conditions will represent the relative contribution of P loads in each of the runoff components, as well as the influence of any point source of P. A.

INFLUENCE OF POINT SOURCESON PHOSPHORUS IN STREAMS

Estimates of the contribution of P to surface waters from domestic wastes in the United States range from 91 x loo to 227 x l o G kg per year with total P concentrations ranging from 3.5 to 9.0 pg/ml (McCarty, 1967; Ferguson, 1968). Weibel et al. ( 1964) estimated that P discharged as raw sewage from combined storm sewers in Cincinnati, Ohio, amounted to 3.4 kg/ha per year as total dissolved P. In the area of Madison, Wisconsin, the per capita contribution of P to surface waters from treated domestic waste was estimated to be 0.544 kg/capita per year (Sawyer, 1947), whereas an estimate of 1 kg/capita per year was given by Metzler et al. (1958) for Chanute, Kansas. The difference between the estimates of Sawyer (1947) and Metzler et al. (1958) may reflect the increased use of P in domestic products, particularly detergents. The impact of sewage outfall on the dissolved inorganic P concentration of streams and rivers was studied by Brink and Gustafsson (1970). Their results are summarized in Table 11. Obviously the impact of the outfall is dependent on factors which include flow rate of the receiving stream and the P content of the effluent. Under certain agricultural management conditions animal excrement may constitute a point source of P to streams. Excrement may enter streams during surface runoff from feedlot operations or by the cleaning of milking sheds into open drains. The magnitude of these sources of P will be discussed later. McCarty (1967) was unable to estimate the magnitude of contributions of P made to streams from industrial wastes. The amounts of P discharged

22

J. C. RYDEN, J. K. SYERS, AND R. F . HARRIS

TABLE I1 Effect of Sewage Outfall on tllc Dissolved Inorganic Phosphorus Concentration of the Receiving Water" IXssolved inorganic P concentration

(pg

P/ml)

Receiving water

Before outfall

After outfall

River Stream Stream Ditch

0.09 0.05 0.11 0.01

0 4% 0.18 4.30 0.75

Data from Brink and Gustafsson (1970).

to streams will depend on the industrial process concerned and local legislation covering the discharge of industrial effluent. Mackenthun et al. (1968), for example, estimated that a potato canning factory and a woollen mill contributed 3446 and 835 kg of P per year, respectively, to the East Branch of the Sebasticook River, Maine. Domestic and many industrial wastes not only supply large amounts of total P to streams but also have a pronounced effect on the concentrations of dissolved forms of P in the receiving stream. Because domestic and industrial wastes are point sources, they are easily recognized within a watershed and are amenable to direct manipulation.

B. RUNOFFFROM FORESTWATERSHEDS A compilation of data from several studies of the quantities of P lost in streanis from stable forest and woodland watersheds is presented in Table 111. Exports of P in streams from long-established and stable forest watersheds provide a useful datum line against which losses of P from other land-use areas may be compared. The data in Table I11 show a considerable degree of uniformity. Total P losses range from 0.68 to 0.02 kg/ha per year with three out of the four values being less than or equal to 0.1 kg/ha per year. Only a few measurements have been made of the dissolved inorganic P concentration of stream water in forested watersheds. The values reported by Brink and Gustafson (1970) in Sweden show a mean of 0.015 pg/ml, with this fraction amounting to 33% of the total annual loss of P. The data suggest that total P and dissolved inorganic P concentrations rarely exceed 0.115 and 0.025 pg/ml, respectively. Two interesting points arise from the data in Table 111. From the study of a stream draining a

z

TABLE I11 Losses of P in Streams Draining Forest Watersheds

Study

Location

Bormann et al. (1968) Brink and Gustafsson (1970)

New Hampshire Sweden

Cooper (1969) Jaworslii and Hetting (1970) Sylvester (1961)

N. Minnesota Potomac River Basin Washington

Taylor et al. (1971)

Coshocton, Ohio

Form measured

P loss (kg/ha/yr)

Total P Total P Dissolved inorganic P Not specified Total P Total P Dissolved inorganic P Total soluble P

0.02 0.06 0.02 0.1s 0.1 0.68 0.07 0.05

P concentration in streamwater (pg P/ml) Range

Mean

-

-

0.008-0.053 0.002-0.026 0.043-0.060

0.048 0.015 0.041

0.024-0.115 0.004-0.009

0.069 0.007 0.015

-

0.011-0.OPO

-

El z 8 2 2 0

c.4 q

+

3 v)

+I

b

24

J . C. RYDEN, J . K. SYERS, A N D R. F. HARRIS

woodland area at Coshocton, Ohio, to which no fertilizer P had been applied for over 30 years (Taylor et al., 1971), it would appear that the woodland is conservative of P. The average total soluble P content of rainfall was 0.020 pg/ml, whereas that in the stream draining the watershed was 0.015 pg/ml. The extent of addition of total dissolved P to the woodland can be calculated from precipitation data given by Taylor et al. (1971 ) ; a value of 0.17 kg/ha per year is obtained. This value is more than three times greater than the annual P loss in the stream. The conservative nature of forests for P is further borne out by the fact that the annual contributions of P to the land surface in precipitation, quoted previously, are in most cases considerably greater than annual exports of P in streams from forest watersheds. In many cases there is an order of magnitude difference. This hypothesis assumes that data covering the P content of precipitation are correct. The second point of interest relates to the “background” P concentration in forest streams. The data suggest only minor seasonal fluctuations in P concentrations, particularly that of dissolved inorganic P. As a major portion of streamflow is considered to have a groundwater origin (Biggar and Corey, 1969; Johnson and Moldenhauer, 1970), it is conceivable that the dissolved inorganic P load in streams of forested areas is primarily due to that in groundwater runoff. If the reported mean P concentrations of forest streams are compared to those for groundwaters, a marked similarity is observed. Juday and Birge ( 193 1) found that the total dissolved P concentrations of 19 wells in northern Wisconsin, an extensively forested area, ranged from 0.002 to 0.197 pg/ml, with an average of 0.018 pg/ml when the highest value is omitted. This mean value is, if anything, slightly higher than the mean concentrations for dissolved fractions of P reported in Table 111. The higher mean concentrations of total P probably arise from suspended inorganic and organic solids that enter streamflow due to turbulence, especially during high flow. The minor fluctuations in P concentrations reported for forest streams suggest that P export is minimally affected by P input from surface runoff. Amounts of surface runoff in forest watersheds will be low owing to the protection afforded by canopy cover and/or forest floor vegetation. The “background” P export in forest streams is a direct reflection of the chemical and physical factors that affect P concentrations in groundwater and subsurface runoff. Because larger amounts of stream flow from forest watersheds will arise from groundwater and subsurface runoff, the “chemical sieving” action of the soil plays a major role in maintaining the consistently low dissolved inorganic P concentrations in forest streams and may also account in part for the apparent conservative nature of forest watersheds for P.

PHOSPHORUS IN RUNOFF AND STREAMS

25

C. RUNOFFFROM AGRICULTURAL WATERSHEDS The loss of P in streams draining agricultural (in most cases arable) watersheds is far less well defined than that for forest streams. This is probably due to the fact that in studies designed to estimate this loss, little differentiation has been made with respect to the forms of runoff. Consequently, there are major problems in estimating P loss from agricultural watersheds using many of the data presented in the literature. Losses of P from agricultural land have not only been based on analyses of streams draining a specific watershed (Campbell and Webber, 1969; Taylor et al., 1971 ), but have also been estimated from data obtained in surface runoff studies (Timmons et al., 1968). Many previous reviews of this subject have relied on such data (Taylor, 1967). Losses of P in streams draining various agricultural watersheds are summarized in Table IV. The lowest loss of total P is from rangeland in Ontario, Canada (Campbell and Webber, 1969) which had received no P fertilizer in living memory. This loss is very similar to losses of total P from forest watersheds, suggesting a minimal contribution if P from surface runoff. Similarly, the total P carried in the base flow, primarily attributable to groundwater runoff, of several streams draining arable agricultural watersheds in S.W. Wisconsin (Minshall et al., 1969) is also little different from total P loads in streams draining forest watersheds. Minshall et al. (1969) reported the total P loss in base flow to be less than 0.12 kg/ha per year. If stream flow during periods of surface and subsurface runoff is included, however, the estimated annual loss of total P increases by one order of magnitude, as indicated by the data of Witzel et al. (1 969) for the same area of S.W. Wisconsin (Table I V ) . These studies suggest that the groundwater runoff or base-flow component of streams draining agricultural watersheds is little different from the total P load of forest streams. It is therefore necessary to estimate the extent to which the P load of streams draining agricultural watersheds may be augmented by P loads of surface and subsurface runoff. The major factors affecting the loads of P in surface runoff from agricultural land include time, amount, and intensity of rainfall, rates of infiltration and percolation, slope, soil texture, nature and distribution of native soil P, P fertilization history, cropping practice, crop type, and crop cover density. A selection of reported losses of P in surface runoff from arable land of various slopes and cropping practices is summarized in Table V. Losses range from the extremely high values of 67 kg/ha per year to almost zero. Losses of P in all studies listed in Table V have been based on the collection of surface runoff (water and particulates) from small experimental

TABLE I V Losses of P in Streams Draining Agricultural Watersheds

Study Campbell and Webber (1969) Fippin (1945) Taylor d al. (1971)

Witzel d al.

Location

S. Ontario, Canada Tennessee Valley Coshoeton, Ohio S.W. Wisconsin

(1969)

a

January through September 1967.

Soil texture

Form of P measured

Slope (%)

I-r

P

Crop

P applied (kg/ha/yr)

P lost (kg/ha/yr)

-

Total P

-

90% Rangeland

0

0.08

-

Total P

-

Row crops, open farmland 50% Permanent pasture; 50% winter wheatmeadow 100% Pasture cultivation-haypasture

-

6.26

3.5

0.07

Silt loam

Total dis-. solved P

Silt loam

Total P

12-18

6-8

1.88

4.08

1.000

9.64 4.27

1.51 1.20

I-r ?c

PHOSPHORUS IN RUNOFF AND STREAMS

27

plots frequently no larger than 30 x 6 m, with subsequent analysis for one or more forms of P. Although this approach was originally developed to investigate soil fertility losses due to soil erosion, it is still used to estimate P loads in surface runoff as it relates to the fertility of surface waters (Timmons et al.,1968; Nelson and Romkens, 1969). It is difficult to make any generalizations regarding the P loads carried in surface runoff or to draw conclusions from them in terms of how agricultural practices and natural variables affect P loads in streams draining agricultural watersheds. This is due to the differences in forms of P measured and the lack of comparative studies with respect to slope, soil texture, cropping, and climatic variables. One of the few studies from which meaningful interpretations of P loss in surface runoff can be made in relation to degree of slope and cropping practice is that by Massey et al. (1953) in Wisconsin (Table V ) . As expected, greater “available” (water-soluble plus pH 3 extractable) P losses to surface runoff occurred on the steeper slopes when cropping practice was kept constant. The introduction of two years hay into the rotation reduced the P loss by a factor of approximately four. The value of “improved” or “conservative” agricultural practices in reducing the magnitude of P losses is illustrated in the studies at Coshocton, Ohio (Weidner et al., 1969) and at Lafayette, Indiana (Stoltenberg and White, 1953). It should be noted, however, that although the “improved” practice reduced the total amounts of acid-hydrolyzable P lost in surface runoff at Coshocton, the concentration of this fraction during surface runoff increased from 0.43 to 0.59 pg/ml. Attempts have been made to measure the relative contributions of the aqueous and particulate fractions of surface runoff to the total loss of a measured form of P. In a plot study using simulated rainfall, Nelson and Romkens ( 1969) obtained dissolved inorganic P concentrations of 0.05, 0.30, and 0.50 pg of P per milliliter in the aqueous phase of surface runoff from fallow plots 12 days after 0, 56, and 1 1 2 kg of P per hectare, respectively, had been disked into the soil, with only slight decreases in concentrations up to 75 days after fertilizer application. Although very high artificial rainfall rates were employed (up to 73.5 mm/hr), indications are that high concentrations of dissolved inorganic P may be maintained in surface runoff water. Timmons et al. (1968) determined the distribution of total P loss in surface runoff between the aqueous and particulate phases from plots under natural precipitation. Although these workers did not report P concentrations, losses of total P in the aqueous phase of surface runoff arising from snowmelt far o.utweighed those in the particulate phase. In contrast, total P loss in the aqueous phase varied in most cases between 5 and 40% of the loss in particulates in surface runoff arising from rainfall.

TABLE V Losses of Phosphorus in Surface Runoff from Field Plots Study

Location

Knoblauch et al.

New Jersey

(1942)

Massey et al.

Wisconsin

(1953)

Soil texture

Form of P measured

Slopc

(%I

Sandy loam

Total P

Silt loam

Soluble pH 3 extractable P (available)

3.5

+

3 3

11 11

20

Stoltenberg and White (1953)

Lafayette, Indiana

Silt loam

0.5 M N H 4 F 0.1 N HCI

+

00 0.5

extractable P (“available”)

Thomas et al. (1968)

Tifton, Georgia

Sandy loam

+

0.05 N HCI 0.025 N H ~ S O I

extractable P

3

Crop

+

Vegetables (i) No manure (ii) Manure (iii) Cover crop (iv) Cover crop manure Corn-oats Corn-oats2 yr hay Corn-oats Corn-oats-? yr hay Corn-oats-4 yr hay Oats-5 yr hay Coma Cornh Soybeansa Soybeansb Wheat” Wheat* Meadow“ Meadow* corn Rye-peanuts-rye Rye-corn-oats Oats-rye

+

P applied (kg/ha/yr)

P lost (kg/ha/yr) 40 06 67 07 59 66 49 65 0 91 0 24 2 91

0 73 0 75 0 13 2 86 0 86

3 82 1 93 0 0 0 0 0 0 0 0

84 48 99 74 02 07 05 02

Timmons et al. (1968)

W. Central Minnesota

Loam

Weinder et al. (1969)

Coshocton, Ohio

Silt loam

a

Total P

Total acidhydrolyzable P

6

-

Fallow Corn-continuous Corn-rotation Oats-rotation Hay-rotation Cornc Cornd Wheat" Wheatd

29.1 99.1 30.2

4.82 17.34 4.83 17.34

Prevailing practice: moderate fertilizer levels; liming t o p H 6.0; straight row planting and cultivation.

* Conservation practice: higher fertilizer levels; liming t o p H 6.5; manure before corn; contour planting and cultivation. c

0.2-0.6 0.1-0. 2 0.1 0.0-0.1 0.1-0.3 10.24 3.11 1.33 0.41

Prevailing practice: straight row tillage across slope; low P fertilizer level; alsike-red clover-timothy meadow mixture; liming t o p H 5.4. Improved practice: contour tillage; high P fertilizer level; clover-alfalfa-timothy meadow mixture; liming t o p H 6.8.

0

2

? w

C

30

J. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

These observations are not unexpected because rainfall tends to loosen soil particles by drop impact, facilitating their entry into surface runoff waters. It is apparent that an appreciable dilution of dissolved P may occur when surface runoff augments base flow in streams. Taylor et al. (1971) reported a mean total dissolved P concentration of 0.022 pg/ml in a stream draining an agricultural watershed at Coshocton, Ohio; concentrations never exceeded 0.100 ,g/ml even under conditions of high stream flow when surface runoff was occurring. It is generally considered that P is retained sufficiently strongly by soil particulates that movement out of the soil profile in percolating waters is minimal (Way, 1850; Black, 1970). Subsurface runoff from agricultural land, however, may contain significant concentrations of dissolved inorganic P in relation to those present in surface waters (Table VI) . It should be noted, however, that the data in Table VI represent losses of dissolved inorganic P in tile and irrigation return flow drains. Artificial drainage systems increase the rates of infiltration and percolation, reducing contact times between the soil solution and soil components capable of sorbing inorganic P from solution. Furthermore, tile drains will remove water from surface horizons of the soil profile, diminishing the possibility for contact between percolating waters and more P-deficient subsoil material. Not all the data in Table VI, however, indicate a net loss of P from the soil profile to subsurface runoff. In the Snake River Valley, Idaho, Carter et al. (1971) found that only 30% of the dissolved inorganic P in irrigation water left an irrigation tract by return flow. When the dissolved inorganic P concentration in irrigation water exceeded 0.010-0.020 pg/ml, irrigation decreased the downstream P load, a useful field example of the chemical sieving action of soils. Johnston et al. (1965), however, reported a net loss of 3% at an applied P fertilizer rate of 51.9 kg/ha on irrigated land in the San Joaquin Valley, California. The data in Table VI indicate that a reasonable proportion of P loss to streams draining arable watersheds can be due to subsurface runoff. Although no data are available to compare P loads due to surface and subsurface runoff, Sylvester (1961) reported that total P loss by irrigation return flow in the Yakima Valley, Washington, ranged from 3.8 to 14.3 kg/ha per year, values higher than many reported for surface runoff losses. Under a nonirrigated farming system, Bolton et a!. (1970) observed losses of dissolved inorganic P in tile drain effluent ranging from 0.13 to 0.29 kg/ha per year at a fertilization rate of 28.9 kg of P per hectare per year. It would appear, therefore, that losses of P in subsurface runoff can be similar or even greater than those in surface runoff. Furthermore, subsurface runoff will occur not only during periods of surface runoff, but also when evapotranspiration is less than infiltration.

TABLE VI Losses of Dissolved Inorganic Phosphorus in Subsurface Runoff Dissolved inorganic P concentration (pglml) Study

Location

Bolton et al. (1970)

Ontario, Canada

Brink and Gustafsson (1970) Carter et al. (1971)

Sweden

Cooke and Wlliams (1970)

Soil texture

Drainage system Tile drains

Clay

-

Snake Valley, Idaho

Calcareous silt loam

Irrigation return flow

Woburn, England

Sandy

Tile drains

Voelecker (1874)

a

S. Central Michigan San Joaquin Valley, California Yakima Valley, Washington

Rothamsted, England

No P fertilizer applied. 28.9 kg P applied per hectare per year.

Clay to sandy loams Heavy silty clay Sandy loam

Clay loam

Range

Mean

Corn, oats Alfalfa, bluegrass

0.200-0.170 0,190-0.270 0.045-0.140

0.180" 0 . 210b 0.079

Alfalfa, corn, root crops, pasture Arable and grassland Arable grassland Root crops

0.007-0.23

0.012

0-0.300

0 . 0uo

0-0.700 0-0.750 0.010-0.300

0.440 0.080

0.053-0.230

0.079

-

Tile drains

Sandy drift Erickson and Ellis (1971 Johnston et al. (1965) Sylvester (1961)

Crop

Tile drains and ditches Irrigation return flow drains Surface return flow drains Subsurface return flow drains Tile drains

Cotton, rice, alfalfa

0.072-0.300

Wheat

-

0.161

0.029-0.460

0.182

0.054-0.802

-

32

J. C. RYDEN, J. K. SYERS, AND R. F . HARRIS

D. RUNOFFFROM LANDASSOCIATED WITH ANIMAL REARING Animal excrement is a source of P to surface waters (McCarty, 1967). Little direct information is available on the P load this source imparts to streams. Although some loss of P to subsurface and groundwater runoff can be expected, the two major ways in which animal excrement may enter streams are by surface runoff from land upon which manure has been spread and by surface runoff from feedlot operations. Manure spread on land in certain areas during the winter months is subject to transport in surface runoff waters. The amounts of surface runoff during spring thaws are particularly great owing to the combined effects of snowmelt and rainfall on frozen ground. A study of P loss from land manured during winter was conducted by Midgley and Dunklee (1 945). Manure was applied to study plots in Vermont during winter for a period of 6 years at a rate of 22.5 tonnes/ha. Losses of total dissolved P amounted to 2.1 and 2.5 kg/ha for 20 and 10% slopes, respectively. It was concluded that P losses were little affected by slope but more by the amount of snow. Using the data of Midgley and Dunklee (1945), Lee et al. (1969) estimated that 6810 kg of P is lost in surface runoff to streams in the Lake Mendota watershed, Wisconsin, from agricultural lands on which manure is spread during 5 months of the year when the ground is frozen. This amounted to approximately 60% of the total P losses from rural land in the watershed. Concern has also been shown in countries where large areas of land are used for pasture and high intensity grazing, such as in New Zealand, that dung pats may be carried in surface runoff and contribute significantly to the P loading of streams (Elliott, 1971). The magnitude of this problem is obviously very difficult to estimate, and its control virtually impossible, unless restrictions are placed on the proximity to streams that stock are allowed to graze. Again, there are few data from which the magnitude of P loss to streams via surface runoff from feedlot operations can be estimated. Surface runoff from feedlots could almost be regarded as a point source of P because such operations are highly concentrated and the area occupied is generally insignificant in relation to the area of the region in which they are located. In Nebraska the total area of concentrated feedlots amounts to no more than 5670 ha (Swanson et al., 1970). When it is considered, however, that cattle of 454 kg average weight excrete 7.7 kg of P per year (Millar and Turk, 1955), of which 60-80% may be in an inorganic form (Peperzak et al., 1959) and that 1.5 million cattle may be on feed at any one time, each occupying an area of 37 m? (Swanson et al., 1970), it is probable that the local impact of surface runoff from these operations on the

33

PHOSPHORUS IN RUNOFF AND STREAMS

P status of streams in the area is considerable. The magnitude of this source of P to streams may result in a spread of effects far beyond the immediate vicinity. Data presented by Gilbertson et al. (1970) for the effect of slope and cattle density on the total P losses from unpaved feedlots in Nebraska are presented in Table VII. The greatest total P losses occurred during snowTABLE VII Effect of Slope and Cattle Density on Total P Loss from ITnpaved Fcedlotsn

Slope 9 6

3

Cattle density (in2/animal) 18.6 9.3 18.6 9.3 18.6 9.5

I,

P loss in winter runoff I, (kg/lla)

P loss i n rain-

80.5 469.3 146.4

58.5 34.2 36.6 29.3 34.2 29.3

256.9

78.1 514.5

storm runoff c (kg/W

Data from Gilbertson et al. (1970). January through April, 1969. April through July, 1969.

melt with a large effect of cattle density but only minor effects due to slope. The latter finding is in agreement with that of Midgley and Dunklee (1945), despite the different purpose of the two studies. The concentrations of total P in surface runoff ranged from 6.8 to 753.2 pg/ml during winter and from 13.9 to 46.6 pg/ml in rainstorm surface runoff. These concentrations are extremely high and would be expected to produce a significant change in the total P concentration of receiving streams.

E. URBANRUNOFF The load of P in streams draining urban watersheds which have a negligible contribution of P from point sources will generally be dominated by that carried in surface runoff. Drainage patterns in urban areas are, in most cases, altered so drastically by paving, the installation of storm sewers, and the channelization of water courses, that contributions from subsurface and groundwater runoff are probably small. It is probable that a large proportion of precipitation in an urban area, which in a rural area would con-

34

J . C. RYDEN, J. K. SYERS, AND R. F. HARRIS

tribute to subsurface and groundwater runoff, is intercepted by drains and becomes indistinguishable from urban surface runoff. Because many of the studies of surface runoff from urban areas have been conducted in areas served by combined sewer systems it is difficult to estimate the contribution of surface runoff to the P load carried by streams. It is virtually impossible to separate the component appearing in stormwater outlets due to overloading of sanitary sewers during wet weather flow from that due to urban surface runoff (Weibel, 1969). Several studies, however, have been conducted recently with the sole objective of determining the quality of urban surface runoff. A summary of these data is given in Table VIII. One of the first studies conducted was that by Weibel et al. (1964) in an 1 1 ha residential and light commercial section of Cincinnati, Ohio. The maximum mean total dissolved P concentrations were observed during the summer (0.36-0.39 pg/ml) whereas minimum values were observed in winter (0.16 pg/ml). By far the most extensive study of the quality of urban surface runoff took place in Tulsa, Oklahoma (Avco Corporation, 1970). The proportion of unused land, arterial streets, and industrial land were all found to be important in relation to the mean dissolved inorganic P concentrations observed in monitored storm Sewers. The highest annual load of 8.8 kg/ha was for urban surface runoff from a light industrial area, a large proportion of which was still under development. Other test areas, all except one including residential property, gave rise to dissolved inorganic P losses of 1.1 to 3.3 kg/ha per year. The largest loads of P per impervious area were from districts where tree cover was dense. This is probably due to the leaching of dissolved inorganic P from leaves, discussed previously. As reported by Kluesener (1972) and Harris et al. (1972), for urban watersheds in Madison, Wisconsin, leaching of leaves and seeds, coupled with the considerably reduced infiltration characteristics of urban areas, can be expected to result in high concentrations of dissolved inorganic P in urban surface runoff in the spring and autumn. Storm sewers draining runoff from residential areas into Lake Wingra, Madison, were monitored intensively during snowmelt, and spring, summer, and autumn storm runoff events; samples were taken every 2-5 minutes during peak flow and at longer intervals over the entire length of a storm to enable determination of the frequency of sampling needed to obtain, in conjunction with flow data, a reliable estimate of P input loads (Harris et al., 1972). Dissolved inorganic P generally constituted more than 80% of the total dissolved P in runoff at all times of the year. Although dissolved inorganic P was highest in the autumn (up to 2.4 pg/ml) and spring (up to 2.1 pg/ml), immediately following leaf and seed fall, respectively, the relative input

wz

TABLE VIII Losses of Phosphorus in Surface Runoff from Urban Areas

meable area Study Avco Corporation

Sylvester (1961)

Weihel et al. (1964)

P loss

runoff waters (pg/ml)

X 0

?? 1 vl

(%)

Form measured

(kg P/ha/yr)

Range

Mean

2

Tulsa, Oklahoma

37

2 80

0 54-3 49

1 15

Ez

Durham, North Carolina Seattle, Washington

29

Dissolved inorganic P Total P

3 4

0 15-52 50

0 55

Dissolved inorganic P Total P Total dissolved

-

Trace-0 78

0 08

-

0 01-1 40 0 052-1 452

Location

(1970)

Bryan (1971)

V

P concentration in

Imper-

Cincinnati, Ohio

; Z

37

P

0.92

4

o

21 0 '26

P m

>

5

36

J . C. RYDEN, J . K. SYERS, AND R. F . HARRIS

loading of dissolved inorganic P was greater during the snowmelt period (levels of 0.4 to 0.9 pg/ml) because of the large volumes of water discharged in this period. Particulate inorganic P varied from 10 to 80% of the total particulate P and showed no consistent relationship to time of year. Levels of total particulate P tended to increase with increasing runoff flow. A substantial proportion of this particulate P was of sufficiently high density to settle rapidly out of the biologically active lake surface waters and probably have minimal effect on lake P fertility status. On the other hand, low density runoff particulate P may provide an important reservoir of biologically available P in lake waters, especially in late spring and summer when P-deficient algae and aquatic plants will tend to accelerate release of dissolved P from such suspended runoff particulates. Although total P levels during a specific runoff event tended to be highest during the initial flush, total P load was dictated essentially by flow rather than by fluctuations in P composition (Harris et a!., 1972). If these trends recur for runoff from different land-use areas, limited sampling of runoff from representative flow-gauged storm sewers during periods of high Row, and analysis of these samples for dissolved inorganic P and low-density particulate P should provide valid estimations of the loads of biologically important P components in urban runoff. Another potentially important source of P to urban surface runoff is that associated with eroding soil. During urban development, particularly on the fringes of urban areas, large tracts of land are frequently stripped of vegetation and graded, maximizing the possibility of erosion should surface runoff occur. There are no reported studies of P losses from such development projects, but Guy and Ferguson (1970) cite soil loss from highway construction in a watershed in the Potomac River basin. This averaged 1710 tonnes/km' per year over a three-year period. The amount of a specified form of P lost to flowing waters by such severe erosion will depend to a large extent on previous land use. As extensive construction programs frequently utilize land previously under agricultural management, high P losses can be expected, the distribution of inorganic P between the solid and aqueous phases being primarily determined by the nature of the inorganic particulates and the concentrations of dissolved inorganic P in solution (Ryden et al., 1972a,b). There is reasonable agreement between the estimates of the P loads carried in urban surface runoff. In many situations, however, urban surface runoff probably only amounts to a small percentage of that contributed by municipal and industrial wastes. As the amounts of P discharged to streams from the latter sources are reduced, urban runoff will become a much more significant source of P to receiving streams (Weibel et al., 1964).

PHOSPHORUS IN RUNOFF AND STREAMS

V.

37

Impact of Phosphorus Carried in Streams on Standing Waters

The overall, short-term impact of surface runoff-derived P on standing waters is expected to be high because a large proportion of the average annual discharge of P occurs over only short periods of time during the annual cycle. Bryan (1971) pointed out that in Durham, North Carolina, there was an annual average of 40 day-long surface runoff cycles. Consequently, the major portion of the annual stream loading of P is concentrated into only 40 days, which potentially amplifies its effects by a factor of approximately nine. Furthermore, the impact of the P load carried in streams is expected to be greater during late spring and early summer when aquatic microorganisms are in the potentially active growth phase. The extent of any evaluation of the impact of P carried in streams on standing waters with respect to their biological productivity will also depend on the forms of P measured. If the only form measured is total P, then evaluation may be no more than one segment in a nutrient budget for the body of standing water. This is the easiest approach but one which allows no interpretation of the effects on biological productivity. Even when forms of P relevant to biological productivity are measured, a major unknown factor centers around the effects of mixing as streams enter standing waters. It is reasonable to suggest that temperature differences between stream water and standing water will have some effect on the degree of mixing. Stream water of a lower temperature than surface lake water would be expected to sink below the surface (Twenhofel, 1950), possibly minimizing effects on the photic zone. Such a situation may occur in summer and autumn when tcmperature differences will be the greatest. Furthermore, the effect of overall stream water density arising from sediment concentration, particularly during periods of surface runoff, causes entering water to sink to a lower level (Twenhofel, 1950). Temperature and density effects would tend to contribute P to deep waters and sediments, removing the P load from an immediately usable location. When minimal temperature and density differences exist between streams and standing waters, then a direct dilution effect will probably operate, providing entering waters produce enough turbulence to facilitate mixing. Under such circumstances the biological ‘availability of particulate forms of P will also depend on settling times which ultimately remove them from the photic zone. If the standing waters are in a stratified condition, however, entering stream waters could override the thermocline. In this case dilution would be limited by the amounts of epilimnetic waters. In spring-fed lakes, the primary source of water is groundwater runoff. The amount of water entering a lake from this source is difficult to eval-

38

5. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

uate. The P load of groundwater runoff has traditionally been regarded as minimal (Taylor, 1967; Keup, 1968), and there is considerable evidence to support this contention. Rarely do P concentrations (in most cases dissolved inorganic P ) exceed 0.020-0.030 pg of P per milliliter (Juday and Birge, 1931; White et al., 1963; Mackenthun et al., 1968; Cooke and Williams, 1970). The importance of groundwater runoff as a P source to bodies of standing water has been based on flow rates and P concentrations of land springs and wells. Although it seems unlikely that groundwater runoff itself will contribute significant quantities of P to standing waters, the upward movement of ground water through the sediment may cause redistribution of P within the sediment and even release of dissolved P to the overlying water. The magnitude of this effect will depend on the P status and redox status of the sediment, the nature of the P-retaining components, the nature of groundwater entry (point or diffuse) into the lake, and its amount. The relocation of sediment P to groundwater runoff may be why Millar and Tash (1967) estimated that groundwater runoff or springs contributed 24.9% of the P inflow to Upper Klamath Lake, Oregon. At present it is difficult to estimate the impact of runoff- and streamderived P on standing waters, and such considerations can only be made if the forms of P relevant to biological productivity are measured. Furthermore, the mixing effects that occur as flowing waters enter lakes and reservoirs, as well as the potential of bottom sediments for the P enrichment of groundwater runoff, require further investigation.

VI.

Present Status and Outlook

The preceding discussion of the factors affecting the dynamics and loads of P in runoff and streams reveals various gaps in our knowledge and illustrates the problems in interpreting the data thus far obtained. The first and major difficulty in data interpretation and comparison is the lack of uniformity in the forms of P measured. In many cases this makes comparison between different studies virtually impossible, thereby prohibiting estimations of the relative importance of any particular source. In many studies, particularly those relating to surface runoff from agricultural land, the measurement of total P has been favored. This has led to the concept of nutrient budgets for P, whereby nutrient input and output for an ecosystem are used to determine whether P is lost. This approach is favored by Frink (1967, 1971). If the estimates of P input and output are based on measurements of total P, little information is gained because such deter-

PHOSPHORUS IN RUNOFF AND STREAMS

39

minations override any knowledge of the distribution of P between various forms in runoff and streams, some of which will have a greater or lesser effect on the biological productivity of surface waters. Although relatively few studies have been conducted on the P loads of streams and surface runoff from forest and urban watersheds respectively, there is considerable agreement in the results so far obtained. The situation is quite different for P loads in runoff and streams from agricultural watersheds. Frink (1971) stated that an “average” agricultural watershed with respect to P loss is a “useless fabrication.” It would appear, however, that the major problem arises from the lack of relevant information upon which reliable estimates can be made, a situation which has arisen largely because of an apparent lack of definition of the system being investigated. The use of surface runoff plots to determine losses of P from agricultural watersheds presents several problems. Surface runoff is a spasmodic rather that a continuous phenomenon, its composition at any location being highly heterogeneous and likely to change over short distances because the energy of the aqueous component, and therefore its ability to carry particulate material, varies with slope. The studies cited previously (Timmons et al., 1968; Nelson and Romkens, 1969), in which attempts were made to measure the distribution of the P load between the solid and aqueous phases of surface runoff appear to have limited value. When surface runoff enters streams, a much greater degree of homogeneity will be assumed, resulting in a new and probably more stable distribution of P between the aqueous and sediment phases, as discussed previously. Measurement of dissolved P fractions in surface runoff itself may lead to erroneous conclusions regarding its impact on the dissolved P status of streams due to the transitory nature of surface runoff. In order to obtain more meaningful estimates of P loss from agricultural watersheds, detailed studies of the P load of streams draining the watersheds are required. Some such studies have been conducted (Minshall et al., 1969; Witzel et al., 1969; Campbell and Webber, 1969; Taylor et al., 1971); these will be referred to as watershed analyses herein. None of the watershed analyses cited, however, covered more than a 2-year period of monitoring; the duration of the study could lead to considerable variation in P loss estimates, due to yearly differences in weather patterns as noted by Timmons et al. (1968) for surface runoff studies. Future studies must be based on the watershed analysis approach in order to avoid bias in estimates of the P loss obtained in plot studies due to differences in the energy of surface runoff imparted by slope variations within the watershed. Furthermore, it is essential that studies be long-term to minimize yearly variation in weather patterns and that the forms of P measured be standardized. Although watershed analyses combine the P

40

J . C. RYDEN, J. K. SYERS, A N D R. F. HARRIS

loads of surface, subsurface, and groundwater runoff, these may be separated by determining P loads under various flow conditions in a way similar to that used by Minshall et al. (1969) and to some extent Taylor et al. ( 1971 ) . With careful selection of small watersheds in the same geographic and climatic area, accurate records of fertilizer practice, and cognizance of less diffuse or even point sources of P (e.g., effluent from animal-rearing or industrial operations) within the watershed, it should be possible to obtain meaningful estimates of the effects of various land use and fertilizer practices as well as physical variables on the loss of P from agricultural watersheds. This approach is similar to that which has been used to evaluate P loads in streams draining forest watersheds. It is also important that this be coupled with investigation to define diffuse sources of P more adequately in terms of the components which constitute such sources. Attempts have been made in this direction, as illustrated in the studies conducted by Taylor and Kunishi (1971), Cowen and Lee (1972), and Ryden et al. (1972a,b). Studies similar to these are necessary if any remedial steps are to be taken to reduce the magnitude of man-induced diffuse P sources and will be particularly valuable if carried out in conjunction with watershed analyses. Only by adopting such an approach will it be possible to provide adequate estimates of the potential of soil and fertilizer P for the P enrichment of streams; a topic which is currently surrounded by considerable controversy. Comparative tables of the relative magnitude of various P sources have been drawn up for individual watersheds (Miller and Tash, 1967; Lee et al., 1969; Jaworski and Hetting, 1970). Although such tables are useful for identification of problems within a specific watershed, extrapolation of this concept to a national basis is dangerous. Local and regional variations in land use can seriously distort the relative impact of any source of P on water quality. The way in which P source data are presented can also lead to different conclusions as to the impact of one source as opposed to another. This is particularly true for comparative tables of P sources compiled on a nationwide basis. McCarty (1967) estimates that in the United States, 4.9 X loGto 77.2 X loGkg of P per year is lost to surface waters through urban surface runoff, whereas 54.5 x lo6 to 544.8 x loG kg of P per year originates from agricultural runoff. If losses are expressed on a per area basis, relative contribution estimates are very similar if not reversed, losses being 0.23 to 3.59 and 0.12 to 1.23 kg/ha per year, respectively. These figures show the need for careful evaluation of problems within any given watershed or group of watersheds. Watershed analyses will provide more useful data than estimations of the magnitude of various P sources from a national standpoint.

PHOSPHORUS IN RUNOFF AND STREAMS

41

ACKNOWLEDGMENTS Research supported by the College of Agricultural and Life Sciences, University of Wisconsin, Madison, by the Office of Water Resources Research Project No. WRC 71-10 (OWRR A- 038- WIS), and by the Eastern Deciduous Forest Biome Project, International Biological Program, National Science Foundation subcontract 3351, under Interagency Agreement AG-199, 40-193-69, with the Atomic Energy Commission, Oak Ridge National Laboratory.

REFERENCES Allen, S. E., Carlise, A., White, E. J., and Evans, C. E. 1968. J . Ecol. 56, 497-504. Anderson, G. 1967. 111 ‘‘Soil Biochemistry” (A. D. McLaren and G. H. Peterson, eds.), pp. 67-90. Dekker, New York. Avco Corporation. 1970. “Storm Water Pollution from Urban Land Activity.” U.S. Dept. of the Interior, Fed. Water Qual. Admin., U S . Govt. Printing Office, Washington, D.C. Bache, B. W. 1964. J. Soil Sci. 15, 110-116. Barber, S. A., Walker, J. M., and Vasey, E. H. 1963. Trans. J . Meet. C o m m . IV and V . Int. SOC.Soil Sci., 1962, p p . 121-124. Biggar, J. W., and Corey, R. B. 1969. I n “Eutrophication: Causes, Consequences, Correctives,” pp. 404-445. Nat. Acad. Sci., Washington, D.C. Black, C. A. 1970. In “Agricultural Practices and Water Quality” (T. L. Willrich and G. E. Smith, eds.), pp. 72-93. Iowa State Univ. Press, Ames. Bolton, E. F., Aylesworth, J. W., and Hore, F. R. 1970. Can. J . Soil Sci. 50, 275-279. Bormann, F. H., Likens, G. E., Fisher, D. W., and Pierce, R. S. 1968. Science 159, 882-884. Brink, N., and Gustafsson, A. 1970. Lantbriikslroegsk. Inst. Markvetenskap-Vattenvard No. I. Bromfield, S. M. 1965. A u s f . J . Soil Rcs. 3, 31-44. Bromfield, S. M., 1967. Arcst. J. Soil Res. 5, 93-102. Bryan, E. H . 1971. S. Water Res. Pollilt. Contr. Conf. 20tlr, 1971. Buehrer, T. F., and Williams, J. A. 1936. Ariz., A y r . Exp. Sta., Bull. 64, 1-40. Campbell, F. R., and Webber, L. R. 1969. J . Soil Water Conserv. 24, 139-141. Carter, D. L., Bondurant, J. A,, and Robbins, C. W. 1971. Soil Sci. SOC. Amer., Proc. 35, 331-335. Chakravart, S. N., and Talibudeen, 0. 1962. J . Soil Sci. 13, 231-240. Chang, S. C., and Jackson, M. L. 1957. Soil Sci. 84, 133-144. Cole, C. V., Olsen, S. R., and Scott, C. 0. 1953. Soil Sci. SOC. Amer., Proc. 17, 352-356. Cooke, G. W., and Williams, R. J. B. 1970. J . SOC. Water Treat. Exani. 19, 253-276. Cooper, C. F. 1969. I n “Eutrophication: Causes, Consequences, Correctives,” pp. 446-463. Nat. Acad. Sci., Washington, D.C. Cosgrove, D. J . 1967. I n “Soil Biochemistry” (A. D. McLaren and G. H. Peterson, eds.), pp. 216-228. Dekker, New York. Cowen, W., and Lee, G. F. 1972. Leaves as a source of phosphorus. Mimeo. Water Chem. Program, University of Wisconsin, Madison.

42

J. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

Dion, H. G. 1944. Soil Sci. 58, 41 1-424. Elliott, I. L. 1971. Proc. Tech. Conf. N . Z . Fert. Mfr. Res. Ass., 13th, 1971 p. 93. Erickson, A. E., and Ellis, B. G. 1971. Mich., Agr. E r p . Sta., Bull. 31, 1-16. Ferguson, F. A. 1968. Environ. Sci. Technol. 2, 188-193. Fippin, E. 0. 1945. Soil Sci. 60, 223-239. Frink, C. R. 1967. Environ. Sci. Technol. 1, 425-428. Frink, C. R. 1971. I n t . Symp. Ident. Measure. Environ. Pollut., Sess. B . June 14 1971. Gardner, W. R. 1965. I n “Soil Nitrogen” (W. V. Bartholomew and F. E. Clark, eds.), Agron. Monogr. No. 10, pp. 550-572. Amer. SOC. Agron., Madison, Wisconsin. Gastuche, M. C., Fripiat, J. J., and Sokolski, S . 1963. Pedologie 13, 155-180. Gilbertson, C. B., McCalla, T. M., Ellis, R. J., Cross, 0. E., and Woods, W. R. 1970. Nebr., Agr. Expr Sta., Bull. 508, 1-23. Gorbunov, N. I., Dzydevich, G. S., and Tunik, B. M. 1961. Sov. Soil Sci. 11, 1252-1259. Gore, A. J. P. 1968. J. Ecol. 56, 4 8 3 4 9 5 . Goring, C. A. I., and Bartholomew, W. V. 1950. Soil Sci. SOC. Amer., Proc. 15, 1 89-1 95. Greaves, M. P., and Wilson, M. J. 1969. Soil Biol. & Biochem. 1, 37-43. Greenland, D. J. 1965. Soils Fert. 28, 415-425. Guy, H. P. 1970. In “Techniques of Water-Resources Investigations of the United States Geological Survey,” Book 3, Chapter CI, pp. 1-55. U.S. Dept. of the Interior, U.S. Govt. Printing Office, Washington, D.C. Guy, H. P., and Ferguson, G. E. 1970. J. Soil Water Conserv. 25, 217-221. Hanapel, R. J., Fuller, W. H., and Fox, R. H. 1964. Soil Sci. 97, 421-427. Harris, R. F., Ryden, J. C., and Syers, J. K. 1972. Agron. Abstr. p. 180. Hemwall, I. B. 1957. Advan. Agron. 9, 95-111. Hingston, F. J., Atkinson, R. J., Posner, A. M., and Quirk, J. P. 1969. Trans. Int. Congr. Soil Sci., 9th, 1968 Vol. 1, pp. 669-679. Hsu, P. H. 1964. Soil Sci. SOC. Amer., Proc. 28, 4 7 4 4 7 8 . Huffman, E. 0. 1969. Trans. I n t . Congr. Soil Sci., 9th 1968 Vol. 2, pp. 745-754. Jackson, M. L. 1963. Clays Clay Minerals, Proc. Nut. Conf. Clays Clay Minerals 11, 29-46. Jackson, M. L. 1969. Trans. Int. Congr. Soil Sci., 9th, 1968 Vol. 4, pp. 281-292. Jaworski, N. A., and Hetting, L. J. 1970. I n “Relationship of Agriculture to Soil and Water Pollution,” pp. 134-146. Cornell Univ. Press, Ithaca, New York. Johnson, H. P., and Moldenhauer, W. C. 1970. I n “Agricultural Practices and Water Quality” (T. L. Willrich and G. E. Smith, eds.), pp. 3-20. Iowa State Univ. Press, Ames. Johnston, W. R., Ittihadieh, F., Daum, R. M., and Pillsbury, A. F. 1965. Soil Sci. SOC.Amer., Proc. 29, 287-289. Jordan, D. 0. 1955. In “The Nucleic Acids” (E. Chargaff and J. N. Davidson, eds.), Vol. 1, pp. 447-492. Academic Press, New York. Juday, C., and Birge, E. A. 1931. Wis. Acad. Sci. Arts Lett. 26, 354-382. Kafkafi, U., Posner, A. M., and Quirk, J. P. 1967. Soil Sci. SOC. Amer., Proc. 31, 348-353. Keup, L. E. 1968. Water Res. 2, 373-386. Kittrick, J. A., and Jackson, M. L. 1956. J . Soil Sci. 7 , 81-89. Kluesener, J. 1972. Ph.D. Thesis, University of Wisconsin, Madison.

PHOSPHORUS IN RUNOFF AND STREAMS

43

Knoblauch, H. C., Kolodny, L., and Brill, G. D. 1942. Soil Sci. 53, 369-378. Kunishi, H. M., Taylor, A. W., Heald, W. R., Gburek, W. J., and Weaver, R. N. 1972. 1. Agr. Food Chem. 20, 900-905. Lahav, N., and Bolt, G. H. 1963. Nufure (London) 200, 1342-1344. Langbein, W. B., and Iseri, K. T. 1960. U S . , Geol. Surv., Water-Supply Pup. 1541-A, 1-26. Larsen, J. E., Langston, R., and Warren, G. F. 1958. Soil Sci. SOC. Amer., Proc. 22, 558-560. Latterell, J. J., Holt, R. F., and Timmons, D. R. 1971. I. Soil. Water Conserv. 26, 21-24. Lee, G. F. 1970. “Eutrophication Information Program,” Occ. Pap. No. 1. University of Wisconsin, Madison. Lee, G. F., Beatty, M. T., Corey, R. B., Fruh, E. G., Holt, C. L. R.,Jr., Hunter, W., Lawton, G. W., Peterson, A. E., Schraufnagel, R. H., and Young, K. B. 1969. “Revised Report on the Nutrient Sources to Lake Mendota” Univ. of Wis. Report to Lake Mendota Problems Committee. Lindsay, W. L., and Stephenson, H. F. 1959. Soil Sci. SOC. Amer., Proc. 23, 12-18. McCarty, P. L. 1967. I . Amer. Water Works Ass. 59, 344-366. McKee, G. D., Parrish, L. P., Hirth, C. R., Mackenthun, K. M., and Keup, L. E. 1970. Water Sewage Works 117, 246-250. Machenthun, K. M. 1965. “Nitrogen and Phosphorus in Water.” U.S. Dept. of Health, Education and Welfare, Public Health Serv., Div. of Water Supply and Pollution Control, U.S. Govt. Printing Office, Washington, D.C. Mackenthun, K. M. 1968. I. Amer. Water Works Ass. 60, 1047-1054. Mackenthun, K. M., Keup, L. E., and Stewart, R. K. 1968. I. Water Pollirt. Control Fed. 40, 72-81. McKercher, R. B. 1969. Trans. Int. Congr. Soil Sci., 9th, 1968 Vol. 3, pp. 547-553. Martin, J. K. 1964. N.Z. J . Agr. Res. 7, 736-749. Massey, H. F., and Jackson, M. L. 1952. Soil Sci. SOC. Amer., Proc. 16, 353-356. Massey, H. F., Jackson, M. L., and Hayes, 0. E. 1953. Agron. I. 45, 543-547. Metzler, D. F., Culp, R. L. Staltenberg, H. A,, Woodward, R. L., Walton, G., Chang, S. L., Clarke, N. A., Palmer, C. M., and Middleton, F. M. 1958. 1. Amer. W a f e r Works Ass. 50, 1021-1057. Midgley, A. R., and Dunklee, D. E. 1945. Vt., Agr. E x p . Sta., Bull. 523, 1-19. Millar, C. E., and Turk, L. M. 1955. “Soil Fertility.” Wiley, New York. Millar, W. E., and Tash, J. C., 1967. Interim Report, Upper Klamath Lake Study, Oregon. Fed. Water Pollut. Contr. Admin. Publ. WP-20-8. Miller, R. B. 1961. N.Z. J . Sci. 4, 844-853. Minshall, N. E., Nichols, M. S., and Witzel, S. A. 1969. Water Resow. Res. 5, 706-7 13. Muljadi, D., Posner, A. M., and Quirk, J . P. 1966. I. Soil Sci. 17, 212-247. Neal, 0. R. 1944.1. Amer. SOC. Agron. 36, 601-607. Nelson, D. W., and Romkens, M. J. M, 1969. In “Relationship of Agriculture to Soil and Water Pollution,” pp. 215-225. Cornell Univ. Press, Ithaca, New York. Nykvist, N. 1959. A c f a Oecol. Scand. 10, 190-21 1 . Ohle, W. 1953. Vom Wasser 20, 11-23. Olsen, S. R., and Watanabe, F. S. 1957. Soil Sci. SOC. Amer., Proc. 21, 144-149. Ozanne, P. G . 1963. Trans. 1. Meet. Comm. IV and V , Int. SOC. Soil Sci., 1962 pp. 139-143.

44

J. C. RYDEN, J. K. SYERS, AND R. F. HARRIS

Peperzak, P., Caldwell, A. G., Hunzicker, R. R., and Black, C. A. 1959. Soil Sci. 87, 293-302. Pierre, W. H., and Parker, F. W. 1927. Soil Sci. 24, 119-128. Pinck, L. A., Sherman, M. S., and Allison, F. A. 1941. Soil Sci. 51, 351-365. Rennie, D. A., and McKercher, R. B. 1959. Can. J. Soil Sci. 34, 64-75. Rogers, H. T. 1941. Soil Sci. SOC.Amer., Proc. 6, 263-271. Roth, C. B., Jackson, M. L., and Syers, J . K. 1969. Clays Clay Minera/s, Proc. Nat. C o n f . Clays Clay Minerals 17, 253-264. Russell, E. J., and Prescott, J . A. 1916. J . Agr. Sci. 8, 65-110. Ryden, J. C., Syers, J. K., and Harris, R. F. 1972a. J. Environ. Qua/. 1, 430-434. Ryden, J . C., Syers, J. K., and Harris, R. F. 1972b. J. Eniziron. Qua/. 1, 435-438. Ryden, J. C., Syers, J. K., and Harris, R. F. 1972c. Analyst 97, 903-908. Sagher, A., and Harris, R. F. 1972. Abstr., C o n f . Gt. Lakes Res., I S t h , 1972 p. 193. Saunders, W. M. H. 1965. N.Z. J . Agr. Res. 8, 30-57. Sawyer, C. N. 1947. J. N. E ~ i g l Water . Works Ass. 61, 109-127. Scarseth, G. D., and Chandler, W. V. 1938. J. Amer. SOC. Agron. 30, 361-374. Shen, M. J., and Rich, C. I. 1962. Soil Sci. Soc. Amer., Proc. 26, 33-36. Shipp, R. R., and Matelski, R. P. 1960. Soil Sci. SOC. Arner., Proc. 24, 450-452. Shukla, S . S., Syers, J. K., Williams, J. D. H., Armstrong, D. E., and Harris, R. F. 1971. Soil Sci. SOC. Amer., Proc. 35, 244-249. Sommers, L. E., Harris, R. F., Williams, J. D. H., Armstrong, D. E., and Syers, J. K. 1972. Soil Sci. SOC.Amer., Proc. 36, 51-55. Stewart, K. M., and Rohlich, G. A. 1967. Eutrophication-A Review. Report to the State Water Control Board, California. Stoltenberg, N. L., and White, J. L. 1953. Soil Sci. SOC.Amer., Proc. 17, 406-410. Stumm, W. 1964. U S . , Pub. Health Serv., Publ. 999-WP-15, 299-323. Sullivan, W. T., and Hullinger, D. L. 1969. Trans. 111. State Acad. Sci. 62, 198-217. Swanson, N. P., Mielke, L. N., and Lorrimor, J. C. 1970. I n “Relationship of Agriculture to Soil and Water Pollution,” pp. 226-232. Cornell Univ. Press, Ithaca, New York. Syers, J. K., and Walker, T. W. 1969. J . Soil Sci. 20, 318-324. Syers, J. K., Shah, R., and Walker, T. W. 1969. Soil Sci. 108, 283-289. Syers, J. K., Murdock, J. T., and Williams, J. D. H. 1970. Sod Sci. Plant Anal. 1, 57-62. Syers, J. K., Evans, T. D., Williams, J. D. H., and Murdock, J. T. 1971. Soil Sci. 112, 267-275. Syers, J. K., Smillie, G. W., and Williams, J. D. H. 1972. Soil Sci. Soc. Amer., Proc. 36, 20-25. Syers, J. K., Browman, M. G., Smillie, G. W., and Corey, R. B. 1973. Soil Sci. SOC. Amer., Proc. 37, 358-363. Sylvester, R. 0 . 1961. U S . , Pub. Health Serv., Publ. SEC-TR-W61-3, 80-87. Taylor, A. W. 1967. 3. Soil Water Conscrv. 5, 228-231. Taylor, A. W., and Kunishi, H. M. 1971. J . A g r . Food Chem. 19, 827-831. Taylor, A. W., Edwards, W. M., and Simpson, E. C. 1971. Water Resour. Res. 7, 81-89. Thomas, A. W., Carter, R. L., and Carreker, J. R. 1968. Trans. A S A E ( A m e r . SOC. Agr. Eng.) 11, 677--679 and 682. Tirnmons, D. R., Burwell, R. E., and Holt, R. F. 1968. Minn. Sci. 24, 16-18.

PHOSPHORUS IN RUNOFF AND STREAMS

45

Timmons, D. R., Holt, R. F., and Latterell, J. J. 1970. Water Rasorrr. Res. 6, 1367-1 375. Twenhofel, W. M. 1950. “Principles of Sedimentation,” 2nd ed. McGraw-Hill, New York. Voelcker, A. 1874. J . Roy. Agr. SUC.Engl. [2] 10, 132-165. Vollenweider, R. A. 1968. “Scientific Fundamentals of the Eutrophication of Lakes and Flowing Waters with Particular Reference to Nitrogen and Phosphorus as Factors in Eutrophication.’’ OECD, Paris. Wang, W. C., and Brabec, D. J. 1969. J . Arner. Water Works Ass. 61, 460-464. Way, J . T. 1850. J . Roy Agr. SOC.EngI. 11, 313-379. Weibel, S. R. 1969. Z i t “Eutrophication: Causes, Consequences, Correctives,” pp. 383-403. Nat. Acad. Sci., Washington, D.C. Weibel, S. R., Anderson, R. J., and Woodward, R. L. 1964. J . Water Pollut. Contr. Fed. 36, 914-924. Weibel, S . R., Weinder, R. B., Cohen, J. M., and Christianson, A. G . 1966. J . Amer. Water Works Ass. 58, 1075-1084. Weidner, R. B., Christianson, A. G . , Weibel, S. R., and Robeck, G. G. 1969. J . Water Polliit. Contr. Fed. 41, 377-384. White, D. E., Hem, J . D., and Waring, G. A. 1963. U.S. Geol. Sun.’., Prof. Pap. 440F, Chapter F. White, E. 1972. N . Z . Ecol. SOC.Proc. 19, 163-172. White, R. W., and Beckett, P. H. T. 1964. Plnnr Soil 20, 1-15. Wier, D. R., and Black, C. A. 1968. Soil Sci. SOC.Arner., Proc. 32, 51-55. Williams, E. G . , and Saunders, W. M . H. 1956. J . Soil Sci. 7, 90-108. Williams, E. G., Scott, N. M., and McDonald, M. J. 1958. J . Sci. Food Agr. 9, 551-559. Williams, J . D. H., and Walker, T. W. 1969. Soil Sci. 107, 213-219. Williams, J. D. H., Syers, J. K., and Walker, T. W. 1967. Soil Sci. Soc. Amer., Proc. 31, 736-739. Williams, J. D. H., Syers, J. K., Walker, T. W., and Shah, R. 1969. Agrocliinzica 6, 491-501. Williams, J. D. H., Syers, J . K., Harris, R. F., and Armstrong, D. E. 1971a. Soil Sci. SOC.Anier., Proc. 35, 250-255. Williams, J. D. H., Syers, J. K., Armstrong, D. E., and Harris, R. F. 1971b. Soil Sci. SOC. Arner., Proc. 35, 556-561. Wischmeier, W. H., and Smith, D. D. 1965. U S , ,D r p . Agr., Handb. 282. Witzel, S . A,, Minshall, N., Nichols, M. S., and Wilke, J . 1969. Trans. A S A E ( A m r r . SOC. Agr. Eng.) 12, 338-341.