Processes controlling soil phosphorus release to runoff and implications for agricultural management

Transcription

1 Nutrient Cycling in Agroecosystems 59: , Kluwer Academic Publishers. Printed in the Netherlands. 269 Processes controlling soil phosphorus release to runoff and implications for agricultural management R.W. McDowell 1,2,A.N.Sharpley 2,, L.M. Condron 3,P.M.Haygarth 4 & P.C. Brookes 5 1 Department of Geography University of Cambridge, Downing Place, Cambridge CB2 3EN, UK; 2 USDA-ARS, Pasture Systems and Watershed Management Research Unit, Curtin Road, University Park, Pennsylvania , USA; 3 Soil, Plant & Ecological Sciences Division, P.O. Box 84, Lincoln University, Canterbury, New Zealand; 4 Institute of Grassland and Environmental Research, North Wyke, Okehampton, Devon EX20 2SB, UK; 5 IACR-Rothamsted, Harpenden, Herts AL5 2QJ, UK; Author for correspondence ( ans3@psu.edu) Received 28 June 1999; accepted in revised form 20 July 2000 Key words: agricultural management, agricultural runoff, animal manures, eutrophication, phosphorus, subsurface flow, surface runoff Abstract Phosphorus (P) loss from agricultural land to surface waters is well known as an environmental issue because of the role of P in freshwater eutrophication. Much research has been conducted on the erosion and loss of P in sediments and surface runoff. Recently, P loss in sub-surface runoff via agricultural drainage has been identified as environmentally significant. High soil P levels are considered as a potential source of P loss. However, without favourable hydrological conditions P will not move. In this paper, we review the basis of soil P release into solution and transport in surface and sub-surface runoff. Our objectives are to outline the role of soil P and hydrology in P movement and management practices that can minimize P loss to surface waters. Remedial strategies to reduce the risk of P loss in the short-term are discussed, although it is acknowledged that long-term solutions must focus on achieving a balance between P inputs in fertilizers and feed and P outputs in production systems. Abbreviations: AAP Adgal-available P; B Bu1k density of the soil; BMP Best management practice; CCB Coal combustion by-product; D Depth of interaction; DP Dissolved P; DPS Degree of P saturation; DRP Dissolved reactive P; DUP Dissolved unreactive P; EPC Equilibrium P concentration; ER Enrichment ratio; K α and β, Kinetic constants related to soil P release; MRP Molybdate reactive P; NPPC National Pork Producers Council; P a Soil available P content; P r Average DRP concentrations in surface runoff; PRP Particulate reactive P; PUP Particulate unreactive P; RP Reactive P; SRP Soluble reactive P; t Time taken of surface runoff; TDP Total dissolved P; abbrevtptotal P; V Volume of total surface runoff during a specific event; VSA Variab1e source area; S Surface runoff suspended soil sediment ratio; USEPA United States Environmental Protection Agency Introduction Phosphorus (P) is required to maintain plant growth. Until recently, the majority of the literature has dealt with the agronomic benefits of P applications to soil. Phosphorus was considered fixed and immobile in the soil and any losses to water were considered incidental from an economic standpoint. Furthermore, manure applications have been based mostly on crop nitrogen (N) requirements. This has resulted in the over application of P fertilizers and manures beyond levels sufficient for optimum crop growth (Sharpley et al., 1994). However, raising soil P concentrations also increases the potential transfer of P from soil to solution and eventually surface water. Increased P concentrations in fresh waters can accelerate eutrophication and impair water use for recreation, drinking and industry (Carpenter et al., 1998).

2 270 The movement of P from soil to water is influenced by many factors, such as the type of inputs (fertilizer and manure applied), outputs dependant upon soil type and management and transport processes dependant upon environmental conditions. The amount of P that reaches fresh waters can be simply viewed as a function of how much P is in the soil and soil hydrological conditions that transport P to surface waters. If we are to mitigate P loss from soil to water, we must understand soil P dynamics and how it interacts with soil hydrology. As it is impossible to control natural rainfall and difficult to influence soil hydrology, this review is more concerned with the movement of soil P from soil to solution and how it can be integrated with hydrology to manage P. Terminology and forms of P We use the following hydrological terminology in this review. Catchment (or watershed). Land surface delineating a drainage area from which water is discharged in stream flow. Stream. Water that flows in a natural channel. Runoff. That part of precipitation which ends up in streams or lakes. Surface runoff (or overland flow). That part of rainfall or snowmelt which flows overland to streams or directly to lakes. Subsurface flow (or drainage). That part or rainfall or snowmelt which infiltrates the soil and moves to streams or lakes as ephemeral, shallow, perched or ground water above the water table. In agriculture, the downward movement (or leaching) of subsurface flow can be intercepted by artificial drainage systems (e.g. tile drains). In addition to surface runoff and subsurface flow other contributors to streams and lakes include direct precipitation and groundwater flow which is precipitation that has become part of the ground water. The literature contains much, and at times confusing terminology for P in solution and soil. Clear P terminology for the total P (TP) in solution was described by Haygarth et al. (1998a) and Haygarth and Sharpley (2000). Confusion can arise when P detected by the method of Murphy and Riley (1962) on undigested samples is considered as free orthophosphate or inorganic P. The Mo reaction may also determine loosely-bound inorganic and organic forms of P, reflecting Mo enhanced hydrolysis (Stevens, 1979), or small colloidal material depending upon filtration (Sinaj et al., 1998). Similarly, unreactive P generally considered to represent organic forms of P, may also contain some condensed forms of P, such as polyphosphates (Ron Vaz et al., 1993). An accurate estimate of orthophosphate is by chromatographic separation (Haygarth et al., 1997), however, this is an expensive and time-consuming method. Most confusion can be overcome by carefully defining P forms determined using wet-chemical and filtration methods. For example, Reactive P (RP) is defined as that which can be detected using the method of Murphy and Riley (1962). Some authors are more descriptive and refer to reactive P in solution as molybdate-reactive P (MRP) (Brookes et al., 1997). Haygarth et al. (1997) reported that a significant quantity of MRP was present in fractions larger than 1000 molecular weight in the freely drained leachate (to 135 cm depth) and storm runoff from grassland soils. By filtering a solution, either through a <0.45 µm membrane (dissolved reactive P, DRP) or through filter paper (soluble reactive P, SRP), dissolved reactive forms of P can be separated from particulate reactive P (PRP). However, it must be recognised that dissolved P (DP) is associated with fine colloidal material <0.45 µm (Haygarth et al., 1997). Total P (TP, unfiltered sample) or total dissolved P (TDP, sample filtered <0.45 µm), is determined using the method of Murphy and Riley (1962), following an appropriate digest as discussed by Rowland and Haygarth (1997). The difference between reactive and total DP is referred to as dissolved unreactive P (DUP) in samples filtered to <0.45 µm and particulate unreactive P (PUP) in unfiltered samples. The term algal-available P (AAP) describes inorganic P which is available to plants and algae. Algal-available P contains DRP and a proportion of particulate P (PP), depending upon filtration. Algae can instantaneously utilise DRP, while PP (organic and inorganic) can also be a long-term source of AAP (Sharpley et al., 1994). Iron (Fe) oxide strips have been used to determine the quantity of AAP within the soil and in runoff (Sharpley, 1993). Additional chemical extractants such as NH 4 F and NaOH have been used to estimate AAP, but are questionable because of the extreme extraction ph used (Golterman, 1988). Relating P inputs to outputs Accounting for the cycling and amounts of P in plants

3 271 and soil is essential for the efficient management and cycling of P. The use of P fertilizer in different regions of the world is determined by a combination of factors, including soil chemical and physical properties, the history and intensity of landuse, and the ability of the land user to afford and/or recognise the need for P fertilizer (Sibbesen and Runge-Metzger, 1995). The intensity of P-input into a catchment can reflect the intensity and type of agricultural development (Smith et al., 1995). Once adequate soil P concentrations are reached, optimum yields for crops can be maintained by replenishing that taken off with fertilizer or manure. A soil containing a surplus of soil P in excess of plant requirements can take a long time to lower concentrations below environmentally acceptable levels. Johnston and Poulton (1976) showed soils which had received no farmyard manure since 1901 took 73 years to deplete Olsen P concentrations from 63.9 and 69.2 to 12.4 and 11.9 mg kg 1 in crop off-takes. A maintenance rather than build-up philosophy to soil P fertility may serve to minimize excessive build-up of soil P concentrations and risk of P enrichment in runoff (surface and subsurface). Williams and Haynes (1992) and Haygarth et al. (1998b) attempted to estimate the P balance in pasture systems at field and regional scales. Williams and Haynes (1992) measured P in soil, plants and animals from an irrigated grazing trial in New Zealand. The total gains and losses of P for unfertilized pasture were compared with those of pasture which had received 188 or 376 kg P ha 1 y 1 as superphosphate for 38 years (Table 1). Of the fertilizer P applied, 35% could not be accounted for in produce removed or in the various residual soil and plant pools. They concluded that most of the unaccounted P may have been removed in surface runoff as suspended soil particles and animal manure during flood irrigation, although some leaching of P may have occurred in the shallow stony soil (Udic distrocrept). Haygarth et al. (1998b) found a surplus in an intensive English dairy farm equivalent to 26 kg P ha 1 yr 1, compared to a Scottish hill farm where the surplus was only 0.24 kg P ha 1 yr 1.Inthe dairy farm, a large proportion of the P was imported via feed rather than fertilizer. In the Netherlands, dietary P intake is being manipulated to reduce the amount of P excreted to land. Morse et al. (1992) noted a 23% decrease of P in faeces and a 17% decrease in total P excretion when dietary P intake of dairy cows was decreased from 82 to 60 g day 1. An increase in dietary P intake Table 1. Total losses and gains of phosphorus over 38 years for the 188 and 376 kg superphosphate ha 1 treatments kg P ha 1 taken from Williams and Haynes, 1992) Measurement Wilderness 188 kg P 376 kg P ha 1 y 1 ha 1 y 1 Fertilizer P applied Total soil (0 20 cm depth) Change in soil P a P removed in animal products P present in roots 8 12 P present in ungrazed herbage Total P accounted for % fertilizer P accounted for a not measured. from82to112gday 1 resulted in a 49% increase of P in faeces and a 37% increase in total P excretion. Poultry feed is commonly supplemented with mineral P, thereby increasing the P concentration of the manure. To decrease the P concentration of the manure, enzymes such as phytase may be added to aid in the digestion and uptake of P from the grain. Alternatively, low phytin grains may be used. A combination of these approaches has the potential to reduce P inputs making it possible to balance P inputs with outputs. Soil P and its movement into solution Inorganic P Soil P exists in inorganic and organic forms. Inorganic P is extremely reactive and exists in the lithosphere in over 150 different mineral forms. These minerals vary in their solubility and ability to supply the soil solution with P. Factors which influence P mineral solubility, include ph, the concentration of aluminium (Al), iron (Fe), calcium (Ca) and magnesium (Mg), the behaviour and surface area of soil particles and soil moisture content. In all soils, a fraction of the P is adsorbed to clay surfaces, especially Al and Fe hydrous oxides and organic matter complexes. Lindsay (1979) suggested that below a ph of 5.8, Al and Fe phosphates control P concentration in soil solution and above this value Ca and Mg phosphates. However, Al, Fe, Ca and Mg phosphates can co-exist over a wide range of soil ph. Due to the low solubility of P minerals the amount of P in soil solution is generally very low. As the soil solution is depleted of P by plant uptake or lost in

4 272 runoff it must be replenished via the solid soil phase. This is controlled by the equilibrium between the soil adsorption system, soil solution and precipitated P compounds (Sample et al., 1980). When a soil is supplied with P, the soils adsorbing constituents becomes increasingly saturated until a point is reached when precipitation of a sparingly soluble compound occurs. The solubility of this compound then determines the upper limit of P concentration. Conversely if the P concentration is lowered, sparingly soluble P will dissolve until the adsorption complex has been saturated to a degree which corresponds to the solubiity of the least stable P compound present (Larsen, 1967). Lookman et al. (1995) showed the existence of two pools of desorbable P in an acidic sandy soil from Belgium, defined according to their kinetics (fast and slow). The fast desorbable P pool, which immediately supplies soil solution, was attributed to the presence of a readily soluble Ca-P (not condensed) or loosely adsorbed (protonated) P which accumulated once a certain degree of P saturation and/or precipitation of Al and Fe oxides had been reached. Using 31 P magic angle spinning nuclear magnetic resonance, Lookman et al. (1997) showed that the amount of the Ca-P compound in the soil decreased after several water extractions. Fardeau and co-authors (for review see Fardeau, 1996) have shown with isotopic exchange kinetics that soil P in many soils or many origins is physically distributed in different compartments which can be more or less quickly exchangeable with DRP in solution. Sinaj et al. (1992) found that P buffering capacity, as determined by isotopic exchange kinetics, was strongly related to the Al and Fe content of a wide range of soils from Albania, and hence could be predicted from the soil parent material. Organic P Organic P forms include relatively labile phospholipids, nucleic acids, inositols, fulvic acids and humic acids. Even though inorganic P has generally been considered the major source of plant available P in soils, the incorporation of fertilizer P into soil organic P (McLaughlin et al., 1988) and lack of crop response to fertilizer P due to organic P mineralization (Doerge and Gardner, 1978), emphasize the importance of organic P in soil P cycling. Sharpley (1985a) found organic P mineralization (15 to 33 kg P ha 1 yr 1 ) in several Oklahoma soils was not completely inhibited by fertilizer P application (20 to 28 kg P ha 1 yr 1 ), with similar amounts of P contributed by both sources. Tate et al. (1991) also found labile organic P mineralization was an important source of P to pasture in both low- and high-p fertility soils in New Zealand. Amounts of P mineralized range from 5 to 20 kg P ha 1 yr 1 in temperate soils and from 67 to 157 kg P ha 1 yr 1 in the tropics, where distinct wet and dry seasons and higher soil temperatures enhance microbial activity (Stewart and Sharpley, 1987). The cycling of P through microbial biomass pools can also be an important process determining organic P dynamics and soil P availability. For example, Brookes et al. (1984) measured annual P fluxes of 5 and 23 kg P ha 1 yr 1 in soils under continuous wheat and permanent grass, respectively, in England. Although biomass P flux under continuous wheat was less than P uptake by the crop (20 kg P ha 1 yr 1 ), annual P flux in the grassland soils was much greater than P uptake by the grass (12 kg P ha 1 yr 1 ). Soil P availability The terms available P and P availability are established in agronomy and used to describe the total plant available P pool and P immediately useable by the plant (Holford, 1997). These terms also describe the quantity of P supplying soil solution (P quantity) and the instantaneous P concentration in soil solution (P intensity). These two factors are linked by the soil s sorptivity or buffer capacity. The capability of the quantity factor to supply the intensity factor is inversely related to the buffering capacity of the soil. The soil s buffer capacity is itself a function of sorption capacity and sorption strength and will control the rate of desorption, and diffusion (Holford and Mattingly, 1976). The rate of dissolution, desorption and diffusion control the kinetics of P release from soil to solution and can collectively be modelled as one process by many kinetic equations, such as the Elovich equation (Polyzopolous et al., 1986; Raven and Hossner, 1994). A soil with a high sorptive power will tend to release P slowly to the soil solution and vice versa, but clearly it is the quantity of P in the soil that ultimately affects P intensity. Dutch scientists extracting soil with acid ammonium oxalate, found the molar concentration of P was related to the average concentration of Al and Fe on a molar basis, and multiplied by 100, gave a degree of P saturation (DPS) (Breeuwsma et al., 1995). Above a DPS of 25% is seen as sufficient for P loss and is used to determine if manure or fertilizer application should occur.

5 273 A large number of chemical solutions exist which aim to measure the quantity of P supplying the soil solution. This reservoir, also known as the labile P pool (Schofield, 1955), consists of adsorbed P, sparingly soluble P compounds, inorganic P in plant residues and some organic P forms. One such extractant, Olsen P (Olsen et al., 1954) used widely in the United Kingdom and New Zealand, extracts a fraction of P adsorbed to Al and Fe oxides, Ca carbonates and Ca-P minerals (Shoenau and Karamanos, 1993). Phosphorus intensity can be directly measured in the soil solution or estimated by extraction of the soil with a dilute salt solution such as 0.01 M CaCl 2 (Schofield, 1955; Aslyng, 1964). Buffer capacity of a soil can be estimated from the initial slope of a P sorption isotherm, which relates the quantity of P sorbed from a solution containing different amounts of P to the P intensity or P concentration remaining in solution. Several workers have shown that P adsorption increases with factors such as temperature, increasing ionic strength and smaller soil to solution ratios and decreases with increasing sodicity of the contacting solution (Barrow, 1980; El Mahi and Mustafa, 1980). A sorption isotherm is also known as a quantityintensity (Q/I) plot. Brookes et al. (1997) generated a Q/I plot from Olsen P and dilute 0.01 M CaCl 2 extracts from plots of a field soil that had a wide range of soil P concentrations (from 10 to 110 mg Olsen extractable Pkg 1 soil). The authors suggested that these extractions could form the basis of an indicator to predict at what concentration of Olsen P above which P movement to soil solution and eventually drainage waters is likely to occur. A similar approach was advocated by Beauchemin et al. (1996) who stated that... a measure of soil capacity of P accumulation (e.g. % DPS) must be accompanied by a measure of desorbability (e.g. water soluble P) to fully assess the risk of contamination of drainage waters by P leaching. The authors also note that the critical level of % DPS above which P is likely to be lost to soil solution and eventually runoff, may vary between soils. The same will occur for a critical level of Olsen P and consequently there is a need to predict this critical value in different soils if it is to be adopted by regulatory authorities. The downward movement of P and subsurface flow For P to be lost in subsurface flow it first must move downwards (defined as leaching) below the soil surface either by matrix flow or by preferential flow in soil cracks, large pores vacated by dead plant roots or earthworm (e.g., Lumbricus terrestris) burrows. Until recently the loss of P by subsurface flow has been considered minimal from agricultural standpoint, but may under certain conditions constitute a significant loss of P from agricultural catchments in terms of the P enrichment of surface waters. Ryden et al. (1973) noted that these losses can be equal if not greater than P lost by surface runoff. The authors cited seven studies with a mean dissolved inorganic P concentration greater that those usually associated with eutrophication of surface waters (from 0.01 mg l 1 [Vollenweider, 1968] to 0.05 mg l 1 [USEPA, 1976]). Sims et al. (1998) cited ten cases where mean orthophosphate (DRP) from artifically drained land was greater than 0.01 mg P l 1. However, in general the P concentrations of water moving through the soil matrix is less than for surface runoff and will decrease as the degree of soil-water contact increases, due to sorption of P by subsoils. Many authors have shown that P has moved (leached) down the soil profile (e.g. Sommers et al., 1979; Lucero et al., 1995). Phosphorus is more susceptible to leaching in sandy soils, soils with a low P retention capacity, waterlogged soils where P is mobilised under reducing conditions and promoted by the application of P inputs (e.g. fertilizers and manure) above plant requirements. Ozanne et al. (1961) reported that 81% of superphosphate applied to a sandy soil in western Australia was leached below the plant roots (> 12 cm). Spencer (1957) measured adsorbed P (NH 4 F extractable) from 45 to 60, 60 to 90 and 90 to 120 cm depths of a Lakeland fine sand in a Florida citrus grove receiving regular application of P fertilizer. Concentrations were approximately 100, 60 and 40 mg kg 1 in limed soils and 150, 100 and 60 mg kg 1 in unlimed soils. Humphreys and Pritchett (1971) measured the distribution of P in six soils from a slash pine operation in Florida, which had received superphosphate for a period of 6 to 10 years. They concluded from measurements of total P that all the fertilizer applied had leached below a depth of 50 cm in two sandy soils (> 93% sand) that had low extractable Al concentrations (<0.10 cmol kg 1 ). Phosphorus leaching can also occur in organic soils and coarse textured mineral soils with high organic matter contents where organic matter accelerates the leaching of P when associated with Fe and Al. Coger and Duxbury (1984) studied P leaching from cultivated organic soils of the 0 to 10 cm depth. They noted

6 274 that the total mineral content of the soils or length of period of cultivation, Ca content and lime applications, or levels of Fe and Al could affect P retention. The authors calculated a P balance and found that 43% of total P in a Carlisle muck soil and 1% of total P in a Palms muck soil could be leached. The large leaching loss from the Carlisle muck soil was attributed to a lower Morgan s extractable Al and Fe content (0.84%) compared to the Palms muck soil (1.9%). In addition to excessive fertilizer application, the spreading of municipal wastewater, domestic sewage, and swine, dairy, beef and poultry manure on land in excessive amounts has been shown to cause P leaching (e.g., Sommers et al., 1979; Nair et al., 1998). Nair et al. (1998) showed that the sorption capacity of the A horizon of several fine sand soils used for dairying in the Lake Okeechobee catchment had been reduced to almost zero. The authors concluded that high P concentrations in the A horizon caused an increase in P concentrations in solutions equilibrated with lower Bh and Bw horizons. Waste application to land is a problem, since unlike fertilizers, the P content of soil and the concentration and form of P in the waste influence the land available for spreading. Furthermore, the P content of manures is seldom considered because spreading manure as a source of nutrients is commonly based on their nitrogen (N) content. Most P leached from sandy soils after fertilizer application is inorganic P, more is as organic P in soils that have a higher organic matter content (Magid et al., 1996). Lucero et al. (1995) found P had leached 15 cm from the soil surface after the application of 28.7 Mt ha 1 of poultry litter. The leaching of P was attributed to mobile organic-p forms, since the clay loam soil had a high P fixation capacity. Frossard et al. (1989) demonstrated some organic P species (adenosine triphosphate, choline phosphate and glucose-6- phosphate) were less strongly sorbed onto soil colloids than inorganic P. Schoenau and Bettany (1987) noted large concentrations of organic P (mainly associated with fulvic acids) in the subsoil of a forested Cryoboralf. The association of P with dissolved organic carbon, either in chemical combination or by electrostatic adsorption could therefore increase P mobility in the profile. The transport of environmentally significant levels of P by subsurface flow is enhanced if the soil is artifically drained. Increased subsurface flow of P may decrease the P lost by surface runoff Bengston et al. (1988) showed that TP in surface runoff was decreased by about 36% over 6 years from tile drained plots (5.0 kg ha 1 y 1 ) compared to undrained plots (7.8 kg ha 1 y 1 ). Haygarth et al. (1998a) showed that the presence of artificial land drains reduced total P export by about 30% (from 3 to 2 kg TP ha 1 y 1 ). This was thought to occur because diverting water flow through deep drains increased contact time with soil surfaces, thus increasing the chance for P sorption, whereas in the absence of drains TP was associated with overland and near surface interflows. Heckrath et al. (1995) demonstrated the movement of DRP, dissolved organic P (DOP) and total particulate P in a heavy clay soil which was neither saturated with P nor had a low P sorption capacity. The authors were able to relate soil test P to P loss in drainage. They noted a distinct point (at about 60 mg kg 1 0.5:M NaHCO 3 extractable P, or Olsen P) above which DRP and TP increased linearly with Olsen P. Sharpley et al. (1977) showed a positive relationship between the TP load of drainage water and the amount of 1.0 M NaCl extractable P from soil at 40 to 50 cm depth. Similarly, Cogger and Duxbury (1984) and Culley et al. (1983) found DRP concentrations from tile drainage were related to the P sorption-desorption properties of peat in New York and Haploquolls in Michigan respectively. By establishing that P losses in drainage waters are related to soil test P, an indicator for the potential loss of P may be developed. However, the frequency, initiation of P loss and to a certain extent the fractionation of P forms (inorganic versus organic and dissolved versus particulate) is determined by hydrology and the underlying hydrological pathways. Stamm et al. (1998) showed that in a drained grassland soil there was an almost immediate response of SRP in tile effluent to manure application and concluded that P was transported through preferential flow paths extending from close to the surface to the drains. Furthermore, in comparison to flow through the soil matrix, soil in long established preferential flow pathways may be P saturated more than the soil matrix yielding a greater potential to transport high P concentrations in drainage water (Jensen, 1998). During transport in subsurface flow P concentrations and forms may change. Haygarth et al. (1998a) showed that the concentrations of Olsen-P were typically larger at the soil surface under permanent pasture than at depth, with a reduction from 18 to 12 mg kg 1 (between only and cm depth) to 7 mg kg 1 at 2 cm depth. This steady decline in the amount of P forms was associated with a change in P forms in different transport pathways. For example, inorganic P was the dominant form in surface pathways (inter

7 275 and overland flow), but the organic forms of P constituted up to 50% of the transfer through deep drains (85 cm deep, Haygarth et al., 1998b). Chardon et al. (1997) have also reported the presence of DOP, which accounted for 70% of TP at cm depth, suggesting a decrease in the proportion of DRP, presumably because of sorption in upper horizons. Therefore, inorganic P reacts with surfaces in upper horizons, but organic forms of P can move more freely. Dils and Heathwaite (1996) showed that the forms of P lost changed from autumn to spring. The largest fraction of total P lost in early autumn was SRP, later in the year storms produced more P from particulate forms and surface runoff. Seasonal changes in soil inorganic P concentrations under various vegetation types are well known. Such seasonal variation could affect the timing of soil sampling to assess fertilizer requirements, the amount of P available for plant uptake and the loss of P in runoff (surface or subsurface). Magid and Nielsen (1992) proposed that soil physiochemical changes associated with soil moisture may mask any biological cycling of soil fauna. Several authors have noted a winter minimum and summer maximum in P concentrations in soil extractants (e.g. bicarbonate) occur in coarse textured soils where a large part of the soil pore space dries out during the summer (Weaver et al., 1988; Magid and Nielsen, 1992). A winter maximum and summer minimum of P in the soil solution have been noted in fine textured soils, where P concentrations may be controlled by the reduction and release of P from ferric hydroxides during wet months (Jensen et al., 1998). Sharpley (1985a) showed that concentrations of organic P and Bray-I P were greatest in winter in a P-fertilized clay soil under a mixed landuse (60% grassland, 40% 3-yr rotation of cotton, oats and sorghum). Cultivation may also affect the amount of P lost by leaching and drainage by disruption of preferential flow pathways. In general, cultivation increases the mineralization of organic P (Gaynor and Findlay, 1995). The surface runoff of P Physico-Chemical pocesses The transport of P in surface runoff as with subsurface flow occurs in dissolved and particulate forms. Particulate P includes P sorbed to soil particles and organic matter that are eroded in surface runoff. This accounts for a large proportion of P transported from cultivated land (75 90%). Surface runoff from grass or forest land carries little sediment, and is therefore dominated by DP (mainly DRP). Loss of P from the soil surface to the stream is controlled by the interaction of source factors (functions of soil, crop and management) with transport factors (surface runoff, erosion and channel processes). Smaller particles, containing a greater concentration of P than large particles, are selectively eroded during surface runoff. The natural log of the ratio of P in suspended sediments to the surface soil, termed the enrichment ratio (ER), is related to the natural log of sediment discharge (Sharpley, 1985b). The transport of DRP in surface runoff and subsurface flow begins with the desorption, dissolution and extraction of P from soil and plant material. In surface runoff, these processes occur as rainfall interacts with a thin layer of surface soil (1 5 cm). Although difficult to quantify in the field, this depth, termed the effective depth of interaction (D), is expected to change relative to rainfall intensity, soil tillage and plant cover (Smith et al., 1993). The DRP in surface runoff from D can be described by a soil P desorption equation: P r = (KP a DBt α S β )/V where P r is the average DRP concentrations in surface runoff (mg l 1 ), K, α and β are constants related to soil surface, clay and organic C content, P a is the soil available P content (mg kg 1 ), B is the soil surface bulk density (Mg m 3 ), t is the duration of surface runoff (min), S is the surface runoff to suspended soil sediment ratio and V is the total surface runoff during a specific event (L). Once in solution the transition of P between dissolved and particulate forms in stream flow is determined by the equilibrium P concentration (EPC, at which no net sorption or desorption occurs) of the contacting suspended sediments, stream bank and bottom sediments and the rate of stream flow. In reality the EPC dictates that an increase in DRP concentration above the EPC results in P sorption by the contacting material. The opposite would occur if DP dropped below EPC. Schreiber (1988) showed that EPC values are highly correlated with extractactable Plevels(0.5M NaHCO 3 and ion exchange resins). There is evidence to show that the kinetics of transition may be masked by stream flow (Gburek and Heald, 1974). The loss of P in surface runoff is promoted by fertilizer or manure application and may be affected by rate, time, method of application, form of fertilizer, amount of time after application, rainfall, slope, tem-

8 276 Figure 1. Relationship between iron-oxide strip P and P sorption saturation of several Oklahoma soils and the algal-available P concentration of runoff (data adapted from Sharpley, 1995). The values on the lines are the regression coefficients of the correlated linear functions obtained for the values displayed. perature, soil type, tillage and vegetation (Table 2). The P lost in surface runoff is generally estimated as less than 5% of that applied (Sharpley et al., 1995a). A number of studies have shown that DRP in surface runoff is a function of soil test P, but the relationship varies with soil type and management. Sharpley (1995) showed that Fe-oxide strip P was linearly related to AAP concentration in surface runoff from 10 soils ranging from sandy loams to clays (Figure 1). Slopes for the regression of DRP against soil test P are generally lower for grass than for cultivated land. However, the range of slopes, means that no one slope can be used for all soils. The principles of DRP release in subsurface flow also apply to surface runoff. Consequently, for a given soil test P concentration, the concentration of DRP in surface runoff will be influenced by P buffering capacity, dependant upon clay, Fe and Al oxides, Ca and Mg carbonates and organic matter. However, contact times between soil and solution in surface runoff are generally much shorter. Sharpley (1995) showed that a clay soil of 200 mg kg 1 Fe-oxide strip P gave a AAP concentration of 0.53 mg l 1 in surface runoff, whereas a sandy loam soil containing a similar concentration of Fe-oxide strip P gave a concentration of 1.65 mg l 1 (Figure 1). When the P sorption saturation of these soils, was calculated using Fe-oxide strip P as extractable soil P one linear regression was able to describe DRP concentrations in surface runoff (Figure 1). Conventional tillage can induce erosion and the potential for P movement. Lindstrom et al. (1992), estimated that the influence of continual ploughing over 100 years (9 slope) had redistributed 25 t ha 1 y 1 of soil, including 2 t ha 1 y 1 of soil P from convex hilltops to concave gullies. Continued tillage, levelling concave gullies (which had previously retained surface runoff) would induce further erosion, while the loss of organic matter-rich topsoil from hilltops reduces soil water holding capacity, induces wind erosion and requires fertilizer inputs to maintain fertility in the remaining soil (Sharpley et al., 1995a). Conservation tillage generally leaves about 30% of the crop residue on the soil surface and has been adopted by many farmers to minimize land deterioration. Leaving a crop residue on the soil surface dissipates rainfall energy, increasing infiltration and thereby reducing runoff and the transport of aggregates (and PP). Andraski et al. (1985) showed that AAP concentrations in surface runoff were less from corn in no-till than from conventionally tilled plots (1.35 m 2 ). Concentrations of DRP and PP in surface runoff from 20 fertilized and unfertilized grassed and cropped plots over a 5-year period decreased when soils were tilled conservatively or not at all (Sharpley et al., 1992) (Figure 2). However, the proportion of AAP to TP in runoff was greater in grassed plots compared to those that were cropped (Figure 2). Several workers have shown a relationship between the rate and method of P application as fertilizer or manure and P losses in runoff. Large amounts of P may be lost when manure or fertilizer application is followed by a period of intense rainfall. Burwell et al. (1975) noted that P in runoff was highest during

9 277 Table 2. Effect of mineral fertilizer and manure application on P loss in surface runoff and fertilizer application on P loss in tile drainage Land use P added Phosphorus loss Percents a Reference and location Dissolved Total applied kg ha 1 yr 1 Surface runoff Mineral fertilizer Grass McColl et al., 1977; New Zealand No-till corn McDowell and McGregor, 1984; Mississippi Conventional corn Wheat Nicolaichuk and Read, 1978; Saskatchewan, Canada Grass Sharpley and Syers, 1976; New Zealand Grass Uhle, 1988; Norway Dairy Manure b Alfalfa Young and Mutchler, 1976; spring Minnesota autumn Corn spring autumn Poultry Manure Grass Edwards and Daniel, 1992; Arkansas Grass Westerman et al., 1983; North Carolina Swine Manure Fescue Edwards and Daniel, 1993; Arkansas Artificial Drainage Corn Culley et al., 1983; Ontario, Canada Oats Potatoes + Wheat Catt et al., 1997; + Barley Woburn, England Minimal till Conventional till Alfalfa Grass 0 30 cm Heathwaite et al., 1997; cm Devon, U.K. Grass Sharpley and Syers, 1979; New Zealand a Percent P applied lost in runoff. b Manure applied in either spring or autumn.

10 278 Figure 2. Decreases in mean annual runoff P concentration (a) and the increase in the proportion of algal-available P as a percent of TP in runoff. (b) as a function of increasing soil cover and soil tillage (adapted from Sharpley et al., 1992). the planting season. It is at this time, when storms are likely, that P is applied and there is minimal plant cover. Less P is lost when the interval between application and surface runoff increases allowing for P to be adsorbed subject to the soils rate of reaction. Phosphorus loss can also be minimized if manure is applied below the soil surface. Ross et al. (1979) demonstrated that injecting dairy manure into a silt loam soil under pasture, almost eliminated P in surface runoff. Hydrologic processes The mechanisms controlling soil P dynamics and release to runoff at point or field scales are known better than the hydrologic controls that link spatially variable P sources and sinks to transport processes within a catchment. This information is critical to the development of effective management programs addressing the decrease of P export from agricultural catchments. Hydrologic and chemical factors controlling P export from catchments are dynamic and highly variable both temporally and spatially (Pionke et al., 1996). Increased precipitation to a catchment increases the amount of discharge and the quantity of P lost by accelerating those transformations that occur during stream flow. A few short, intense storms can account for as much as 90% of the annual P export from catchments and from only 5% of the land area. Mid summer storms which account for only 3% of runoff time were responsible for between 75 80% of TP runoff from the South Pine River in south eastern Queensland, Australia (Cosser, 1989). Runoff production in many catchments in humid climates is controlled by the variable source area (VSA) concept of catchment hydrology (Ward, 1984). Source areas vary in time, expanding and contracting rapidly during a storm as a function of precipitation, temperature, soil type, topography, ground water and moisture status over the catchment (Gburek and Pionke, 1995). Surface runoff from these areas is limited by soil water storage rather than infiltration capacity. This situation usually results from high water tables or soil moisture contents in near-stream areas. The boundaries of surface runoff producing areas will be dynamic both within and between rainfall events (Gburek and Pionke, 1995; Zollweg et al., 1995). During a rainfall event, area boundaries will migrate upslope as rainwater input increases. In dry summer months, the surface runoff producing area will be closer to the stream than during wetter winter months, when the boundaries expand away from the stream channel. The location and movement of VSAs of surface runoff within a catchment are also influenced by soil structure, geologic strata, and topography. Fragipans or other layers such as clay pans can determine when and where perched water tables occur. Shale or sandstone strata may also influence soil moisture content and location of saturated zones. For example, water will perch on less permeable layers in the subsurface profile and become evident as surface flow or springs at specific locations in a catchment. Converging topography in vertical or horizontal planes, slope breaks, and hill slope depressions or spurs, also influence variable source area hydrology within catchments. Net precipitation (precipitation evapotranspiration) governs catchment discharge and thus, TP loads to surface waters. This should be taken into account when comparing the load estimates from different regions. It is also one reason why there seems to be more concern with P in humid regions than in arid regions, although the intensity of agriculture can be higher in arid regions. In catchments where surface runoff is limited by infiltration rate rather than soil water storage capacity,

11 279 areas of the catchment can alternate between sources and sinks of surface flow. This again will be a function of soil properties, rainfall intensity and duration and antecedent moisture condition. Thus, consideration of hydrologic controls and variable source areas is critical to a more detailed understanding of P export from agricultural catchments. Implications for agricuitural phosphorus management The preceding information on the processes and factors controlling P transfer from soil, fertilizer, and manure to runoff water has direct implication to the management of agricultural P to reduce P losses from catchments. A better understanding of how, where, and when P can move from a given site will allow more effective targeting of remedial measures to mitigate P export from catchments. Do we understand the mechanisms by which P is transported from the soil to recommend management practices to mitigate P loss? Extensive research has clearly defined the main mechanisms controlling P transport in surface runoff. Until recently, however, there has been much less focus on subsurface mechanisms. As a result, our understanding of the mechanisms and quantities of P transported by subsurface pathways is less well understood than for surface mechanisms. This review describes two mechanisms by which subsurface transport of P may occur: (1) By leaching P down the profile as matrix flow which is in contact with most of the soil and that is eventually intercepted by artifical drainage or subsurface flow flowing laterally into streams. This is of concern in soils with a low P retention or a soil P concentration in excess of plant needs. There is also a concern that organic P, being more mobile than inorganic P, will leach further down the soil profile. However, it is unclear how algal-available forms of organic P are. There would be a need to assess the bioavailability of organic P especially in high organic matter soils, such as muck soils and permanent grassland compared to crops grown on the same soil. (2) By the transport of P in water that bypasses the soil matrix in preferential flow. This is of concern where preferential flow pathways may represent the major source of P leaching, especially if the soil ehich makes up the preferential flow pathway is more P saturated than the soil matrix. In this situation, P movement tends to be a function of surface management and not soil P concentration. However, in tilled soils the absence of preferential flow pathways in the plough layer (0 to 23 cm) implies that P in preferential flow in the subsoil is supplied by matrix flow (Thomas et al., 1997). Management should not only take into consideration those practices that are likely to increase the risk of P loss to runoff, but also to better manage P applications for the more efficient uptake by the crop. In the long-term, better environmental management of soil P requires a multidisciplinary approach that incorporates agronomic management to better balance soil P inputs with the efficiency of crop uptake and the incorporation of soil hydrology to better estimate those areas actually of risk to P loss in surface runoff or subsurface flow. How do we better manage the land application of P? Confined animal production systems are a large source of P in runoff This can be a particular problem in pasture systems, where annual P requirements are much lower than manure P produced. In many cases, specialised feed rations are purchased off-farm and supplemental feed is often cheaper at a local feed mill than growing it on the farm. One major problem for land users is the cost of moving manure to a greater acreage, where it could supplement or even replace mineral fertilizers. The recent trend in the formation of co-operatives that can more cost-effectively compost and compact manure should be encouraged. Neighbouring landusers and private industry are also developing manureprocessing alternatives. Examples of this include centralized storage and distribution networks, regional composting facilities and pelletizing operations that can produce a value-added processed manure for distribution to other areas. The level of farmer involvement could be linked to the number of animals per farm or the quantity of manure produced. Storage of manure will allow more flexibility in timing applications. For example, a plastic sheeting can be a low cost storage option for some solid manures. How may we mitigate P losses? In addition to managing soil P inputs in feed, manure and fertilizers, possibilities exist to decrease P loss in

12 280 subsurface and surface runoff. For subsurface flow this may include the retention of dissolved P by filtration systems at the drain outflow, deep cultivation to disturb preferential flow pathways, changing landuses and the retention of P by riparian zones or wetlands near the receiving water body. For surface runoff, several amendments with some types of waste products that can sorb large amounts of P, have the potential to be used as a short-term measure to minimize the P enrichment of surface runoff. Recent research has shown that the addition of gypsumcontaining coal combustion products (CCBs) can decrease the amount of water soluble soil P, without appreciably reducing plant available P (measured as Mehlich 3; Stout et al., 1998). These by-products accomplish this by shifting P from the readily available soil P pools to less available soil P pools. Employing such a tactic on critical source areas of a catchment has the potential to make large reductions in P export by treating only a small portion of a catchment (Stout et al., 1998). How do we assess site vulnerability to P loss and link soil P to soil hydrology? Data requirements limit the application of models simulating P transfer from terrestrial to aquatic ecosystems by field personnel, such as farm advisors, extension agents, and consultants. Thus, a simple indexing procedure was developed to rank field vulnerability to P loss and to identify agricultural soils and management practices for targeted remediation (Sharpley et al., 1998; Gburek et al., 2000). The P index account for and rank transport (surface runoff, drainage, erosion) and source factors (soil P, rate, timing and method of P added) controlling P loss in subsurface and surface runoff and identify sites where the risk of nutrient movement is expected to be higher than others. These indices have been incorporated into state and national nutrient management planning strategies at catchment scales, to address the impacts of animal feeding operations on water quality, to help identify agricultural areas or practices that have the greatest potential to impair water resources (Lander et al., 1998; Simpson, 1998; USDA-USEPA, 1998). Clearly, strategies to minimize P transfers are most effective if vulnerable areas within a catchment are targeted, rather than implementation of general strategies over a broad area. Several agencies in the US and Ireland, charged with developing farm nutrient management plans are using the P index to target sites for more intensive manure management based on P content and potential for losses in runoff. According the basic framework of the index, soil test P criteria are quantified to meet local agronomic, environmental and economic needs of the region. This remains the most contentious issue and more information relating soil and runoff P, as shown in Figure 1, is needed to establish environmental risk assessment tools that identify the potential for P loss from a given site or landscape. How can we help farmers adopt measures that protect water resources? Perhaps the most critical need is to initiate real and lasting changes in agricultural production by focusing on consumer based programs and education rather than burdening farmers. This will be the most challenging. Farmers are at the bottom of the decision chain with regional and often global economic pressures and legislation or constraints, which farmers have little or no control over, influencing their decisions. Since World War II, greater fertilizer N availability via increased production and reduced cost, along with soybean breeding, improved breeding, specialized feed concentrates, and new production technologies have led to greater animal productivity on a smaller land area. At the same time, the land base available for manure management has declined due to urban development, idled land, and reforestation. As a result, animal farming has changed from land-based to capital or economically-driven systems. Thus, manure production and management issues facing farmers are to a large extent driven by economic factors rather than environmental issues. Clearly, we have to look at new ways of using incentives to help farmers implement BMPs. The challenge is to recognize how social policy and economic factors influence the nutrient-management agenda. Equally important is that everyone is affected by and can contribute to a resolution of P-related concerns. Rather than assume that inappropriate farm management is responsible for today s water quality problems, we must address the underlying causes of the symptoms. These causes are related to marketplace pressures, the breakdown and imbalances in global P cycling, and economic survival of farms. Research is needed to develop programs that encourage farmer performance and stewardship to achieve agreed environmental goals. These programs should focus on pub-