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a School of Environment & Agriculture, University of Western Sydney (Hawkesbury Campus), Locked Bag 1797, Penrith South, DC NSW 1797, Australia [formerly Summit-Quinphos (NZ) Ltd]
b Summit-Quinphos (NZ) Ltd, PO Box 24-020, Auckland, New Zealand
c National Institute of Water & Atmospheric Research Ltd, PO Box 11-115, Hamilton, New Zealand
* Corresponding author (bquin{at}sq.co.nz)
Received for publication September 26, 2003.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONFORMS OF PHOSPHORUS IN...INFLUENCE OF RECENT FERTILIZER...OVERALL COMMENTSDIFFERENCES BETWEEN TYPES OF...MECHANISMS FOR MANAGING EVENT...CONCLUSIONSREFERENCES |
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Abbreviations: DAP, diammonium phosphate • DAPR, direct-application phosphate rock • DCP, dicalcium phosphate • DIP, dissolved inorganic phosphorus (usually equivalent to DRP) • DP, dissolved phosphorus (usually equivalent to TDP) • DRP, dissolved reactive phosphorus • PP, particulate phosphorus • SSP, single superphosphate • TDP, total dissolved phosphorus • TP, total phosphorus • TRP, total reactive phosphorus • TSP, triple superphosphate
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONFORMS OF PHOSPHORUS IN...INFLUENCE OF RECENT FERTILIZER...OVERALL COMMENTSDIFFERENCES BETWEEN TYPES OF...MECHANISMS FOR MANAGING EVENT...CONCLUSIONSREFERENCES |
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Many soils in their natural state are low in readily available P, and require fertilization to achieve maximum possible yields. The use of fertilizers increased greatly over the last century, with global use of phosphate fertilizers increasing from about 873 million tonnes of P in 1913 to a peak of about 16591 million tonnes of P in the late 1980s (Wild, 1988; United Nations Industrial Development Organisation/International Fertilizer Development Center, 1998). It was once thought that P was completely immobile in soil, and thus farmers could be encouraged to increase phosphate fertilizer application without fear that P applied in excess of crop requirements would be lost from the soil profile (Haygarth and Jarvis, 1999). However, it is now known that P accumulated in soils or freshly applied in fertilizer may be lost from soil through leaching and surface runoff. While subsurface pathways can be significant in P transfer to water, especially in soils with low P-retention properties and/or significant preferential flow pathways (e.g., cracking clay soils), it is reasonably well established that in most watersheds, P export occurs mainly in overland flow (Sharpley and Rekolainen, 1997; Sharpley et al., 1999). This review will therefore focus mainly on P transport via this mechanism.
Agronomically, the amounts of P lost per annum in runoff are generally inconsequential. However, from the perspective of water quality, only very small concentrations of P are necessary for a body of water to become eutrophic. For example, default trigger total phosphorus (TP) and dissolved reactive phosphorus (DRP) values in New Zealand, the concentrations below which there is considered to be a low risk that adverse biological effects will occur, are 26 to 33 and 9 to 10 µg P L–1, respectively, for slightly disturbed systems (Australia and New Zealand Environment and Conservation Council, 2000). Threshold levels of DRP for saturation of algal growth are considered in New Zealand to be 15 to 30 µg L–1 (Ministry for the Environment, 2001).
With the identification and reduction of water pollution from many point sources, attention has been turned toward the contribution from diffuse agricultural sources of P, which are now considered the major source of P in most instances (Haygarth, 1997; Sharpley and Rekolainen, 1997). This has led to a proliferation of research on the subject of P losses from agricultural land, particularly in the last decade or so (e.g., reviews by Carpenter et al., 1998; Nash and Halliwell, 1999; Gillingham and Thorrold, 2000; McDowell et al., 2001).
The mechanisms and pathways of P transfer are complex, and while advances in our knowledge have been made, there is still much that is poorly understood or under-investigated (Leinweber et al., 2002; Sharpley et al., 2002). One such area is the loss of P from recently applied fertilizers and manures, which can be especially important in grassland systems where these P sources are surface-applied. These losses may account for a substantial portion of the total P losses from a given area of land, but this aspect of P transfer has received relatively little attention (Leinweber et al., 2002). Distinguishing the relative contribution made by individual sources is necessary in attempting to minimize loss with targeted land management policies (Edwards and Withers, 1998).
In this paper, we review the literature on losses of P in agricultural runoff, from mainly pastoral land, with particular focus on the potential influence of manufactured and mineral fertilizer applications on P transfer, and attempt to put this P loss factor into perspective with other potential sources such as diffuse losses from background soil P and erosion.
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FORMS OF PHOSPHORUS IN RUNOFF |
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TOPABSTRACTINTRODUCTIONFORMS OF PHOSPHORUS IN...INFLUENCE OF RECENT FERTILIZER...OVERALL COMMENTSDIFFERENCES BETWEEN TYPES OF...MECHANISMS FOR MANAGING EVENT...CONCLUSIONSREFERENCES |
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The literature contains differing assumptions about which form of P, dissolved or particulate, is predominant in the various P transfer pathways. This issue is discussed in detail in the following two sections.
Catchment-Scale Studies
Results from many studies at the catchment scale indicate that particulate phosphorus (PP) is the predominant form exported from agricultural land. In a review of New Zealand plot- and catchment-scale research on P in runoff from pasture, Gillingham and Thorrold (2000) reported a range of values from 62 to 91% of TP being in particulate-associated form, with the highest figure reported by the catchment study of Lambert et al. (1985). In a comparison of adjacent pasture, pine, and native forest catchments with porous soils developed from volcanic ash (Typic Vitrandepts), Cooper and Thomsen (1988) estimated TP export coefficients to be 167, 9.5, and 12.0 kg km–2 yr–1, respectively, of which approximately 22, 38, and 14% were as DRP, respectively. In a high-country agricultural catchment in Otago, New Zealand, consisting of predominantly sandy loam soils developed in loess or weathered schists, DRP concentrations in stream flows were found to range from 11 to 33% of TP at various sites (Caruso, 2000), while in a hill-country catchment at Whatawhata Research Centre, in the Waikato region of New Zealand, consisting of predominantly yellow brown earth soils (Ochreptic Hapludults, Umbric Dystochrepts), DRP was estimated to contribute only about 9 to 19% of TP exported from pasture and mixed land (native forest and pasture) (Quinn and Stroud, 2002).
However, in measurements of P content in stream flow rather than actual overland flow, the proportion of dissolved phosphorus (DP) to particulate phosphorus may be confounded by adsorption of DP onto sediment particles during overland flow and/or in the stream channel itself (Sharpley et al., 1981, 2000). Thus, Cooke (1988), in a 16-ha hill-country catchment, with three main soils groups (Typic Dystrochept, Aeric Haplaquept, and Typic Dystrandept), under permanent ryegrass (Lolium perenne L.)–clover (Trifolium spp.) and grazed by sheep and cattle, found that 63% of P in overland flow was in dissolved form, but more than 85% of P exported from the catchment in streamflow was particulate. Cooper and Thomsen (1988) compared baseflow inputs (spring waters) and outputs from the stream channels in the three study catchments and estimated that the DRP exported decreased by 72, 76, and 87% for the pasture, pine, and native forest catchments, respectively. To quote these authors, "There are clearly processes operating within the stream channels that can drastically alter the quantity of dissolved nutrient that is exported from the catchments." Other authors have made similar observations regarding adsorption of dissolved P during transport (Schuman et al., 1973; Sharpley et al., 1993). It should be noted also that all the above studies were conducted in catchments with moderately steep (average 17°) to steep (>30°) topography, and/or soils prone to erosion, and those under pasture were grazed by sheep and cattle all year round and/or at relatively intensive stocking rates (e.g., 13 stock units ha–1 at Whatawhata). This suggests that overland flow from these sites is likely to contain a relatively high load of suspended solids, which may adsorb significant amounts of soluble P.
In contrast, Wilcock et al. (1999), in their lowland stream study in the Toenepi dairy farming catchment, with a mix of poorly and moderately well-draining silt and clay loam soils, also in the Waikato region, found an average specific DRP yield of 0.54 kg ha–1 yr–1, accounting for about 47% of TP. Such an export of DRP from Toenepi was significantly higher than had previously been described for New Zealand catchments (DRP range = 0.04–0.3 kg ha–1 yr–1, TP range = 0.3–1.7 kg ha–1 yr–1) (Wilcock, 1986). Concentrations of TP and DRP in the Toenepi lowland stream were generally lowest in winter, in contrast to the studies of Cooper and Thomsen (1988) and of Quinn and Stroud (2002), and increased in spring and autumn. The seasonal DRP and TP peak concentrations in the Toenepi stream coincided with times when fertilizer was being spread, and when dairy effluent from two-staged oxidation ponds was discharged (Wilcock et al., 1999). This suggests that intensive water quality monitoring is required in many larger-area and/or time-constrained runoff studies, or otherwise that data obtained failed to accurately reflect the effect of specific fertilizer or effluent applications.
In another recent study assessing water quality in lowland streams in dairy farm catchments, Davies-Colley and Nagels (2002) found that median DRP concentrations in pastoral streams in Westland, New Zealand, accounted for 58 to 76% of TP concentrations, while those of pastoral streams in the Waikato area ranged from 50 to 71% of TP concentrations, even higher than that found by Wilcock et al. (1999). Studies on the lowland catchments found much lower suspended solids loading and turbidity in the stream water than was reported in the hill-country catchment studies, reflecting higher soil erosion in hill country. For example, Wilcock et al. (1999) estimated the specific yield of suspended solids in the Toenepi catchment to be 142 kg ha–1 yr–1, compared with losses of up to 3212 kg ha–1 yr–1 from the Whatawhata pasture catchment (Quinn and Stroud, 2002).
It could be implied from these results, then, that PP is the dominant form in runoff from hill-country pastures, while dissolved forms of P dominate in runoff from lowland grassland systems. However, such a view may be oversimplistic, ignoring the potential chemical dynamics of P during the transport phases in overland and stream flow. Even in hill-country situations, much of the P transported may be in dissolved form initially. The relative proportions of DP to TP in surface runoff may be influenced by a large number of factors, including pasture cover, stocking type and rate, soil erodibility, soil texture and particularly clay content, antecedent soil moisture, and many others. This has major significance in the formulation of methods to manage the potential for loss of P from agricultural systems, as discussed later in this review.
Surface Runoff Studies
A number of studies in New Zealand that measured P contents in surface runoff water rather than stream water have also found PP to be the predominant form transported. Lambert et al. (1985) monitored eight 0.1- to 1.5-ha catchments within a sheep and cattle grazing trial at the Ballantrae Hill Country Research Station, near Woodville, New Zealand. Soils that make up this area are described as yellow brown earths, yellow brown earth intergrades to yellow grey earths, and related steepland soils, formed on Tertiary sandstone, siltstone, and mudstone. Average annual losses of dissolved inorganic phosphorus (DIP) ranged from 9 to 16% of the 0.69 to 1.47 kg ha–1 TP exported. In two hill-pasture enclosures on yellow brown earth soils formed on colluvium and alluvium, at Taita Experimental Station, Hutt Valley, New Zealand, McColl and Gibson (1979b) measured a grand mean concentration of P soluble in 0.5 M H2SO4 of 0.79 g m–3, about 33% of the TP concentration. As with the catchment studies mentioned previously, the experiments at Ballantrae and Taita were conducted on moderate to steep hill country. The average slope at Ballantrae ranged from 19 to 25°, while the mean slope near the individual collectors at Taita varied between 11 and 21.5°. Sediment losses from both sites were relatively high, and were related to timing and type of grazing. Average annual sediment losses from Ballantrae ranged from 1107 kg ha–1 under rotationally grazed sheep to 2743 kg ha–1 under rotationally grazed cattle. Export coefficients were not estimated for the Taita experiment, but mean annual total solids concentrations averaged 231 g m–3 before grazing and 924 g m–3 after grazing. The latter figure is significantly higher than the range of mean storm-flow concentrations from several catchment studies reported by Wilcock (1986) of 50 to 460 g m–3, probably relating to the much steeper topography in the Taita catchment.
Despite the findings of studies such as these, it is often stated in the literature that most P exported in runoff from grassland is usually in the dissolved fraction (e.g., Sharpley et al., 1994, 2000; Nash and Murdoch, 1997). Tham (1983) (cited by Nash et al., 2000b) measured P export from a gently sloping annual pasture under sheep grazing. Up to 70% of the TP in runoff consisted of DRP, despite the runoff containing up to 2000 mg L–1 suspended solids. No explanation for this occurrence was given. Greenhill et al. (1983a)(1983b, 1983c) compared P concentrations in runoff from mown perennial pasture plots on very fine sandy clay loam soils, at three sites in Gippsland, southeastern Victoria, Australia, receiving four rates of P fertilizer. Annual mean TP in runoff from the control plots, which received no fertilizer, consisted of up to 91% DRP. Nash and Murdoch (1997) measured P losses in runoff from a fertile dairy pasture on a yellow podzolic soil, at Darnum in West Gippsland. Over seven runoff events (approximately 31–78 d after fertilizer application), the TP loads in the runoff water ranged from 9 to 1762 g, with a range of DRP content of 78 to 94% of TP (average 89%). Over a period of three years, DRP and total dissolved phosphorus (TDP) exported from this site in runoff were found to account for approximately 93 and 96%, respectively, of TP export (Nash et al., 2000b).
Fleming and Cox (1998) measured P losses in runoff from two dairy pasture catchments, on a series of Ferric Eutrophic Brown and Red Chromosols, Haplic Eutrophic Brown Dermosols, and Mottled Eutrophic Yellow Chromosols, with strong texture contrasts between the A (sandy loam) and B (clay) horizons, at Flaxley Agricultural Centre, South Australia, and reported that between 50 to 60% of the TP lost was in dissolved form. In a continuation of this study, the overall proportion of TP exported as DP over three years was 48 and 62% for the east and west catchments, respectively. Dissolved P concentration in runoff water was greater than that of PP in 17 out of 22 runoff events, with two other events where the proportions were approximately equal (Fleming and Cox, 2001a). The proportion of DP to PP appeared to be related to the soil water status, with a greater proportion of the P in PP form in the wettest year, very little PP in the driest year, and an intermediate proportion in the last year, where rainfall was also intermediate compared with the other two years. Similarly, Melland et al. (2001) measured P runoff from four 0.5-ha phalaris [Phalaris aquatica L. cv. Australian]–subterranean clover [Trifolium subterraneum spp. yanninicum (Katzn. et Morley) Zohary and Heller cv. Trikkala] pastures near Hamilton, Victoria, over a 2-yr period, and found the proportion of TP as TDP to range from 46 to 99%, with higher losses coming from the more fertile pastures. Vollmer et al. (2001)(2002) compared P concentrations in runoff from artificial and natural rainfall at two grassland sites near Zurich, Switzerland, with high Olsen P values (69 and 126 mg P kg–1 soil), and found that DRP was the main fraction in all runoff samples analyzed, with an average value of 69 and 64% from the two sites, respectively. These contrasting results in the ratios of DP to PP in runoff from pasture land raise questions about what the driving factors in P transport dynamics in overland flow really are.
Factors Affecting Phosphorus Fractions in Runoff
A partial explanation of the divergence in findings regarding the relative amounts of the different P fractions lost in runoff may be found in studies where measurements have been taken regularly over a lengthy period. Several of these have found that the proportion of DRP to PP being exported from a given area varies significantly with season. Thus, Cooke (1988) found that PP was the dominant form in surface runoff during the winter and spring months, but that DRP became the dominant fraction over summer and autumn. Fleming and Cox (1998) found that soil water status (duration of saturation) appeared to have an effect on the proportion of P exported as DP or PP, with runoff events where PP was the dominant form occurring during winter of the wettest year of the study (1996). Thus, 57% of the P load was as PP in 1996, but the value ranged from 9 to 35% only in the subsequent, significantly drier years (Fleming and Cox, 2001b). Heathwaite and Dils (2000) found that the concentrations of DIP relative to PP and dissolved unreactive P in overland flow increased from the top of a hillslope to its base, and that the proportions of these fractions also varied significantly over the course of an 18-mo period, although DIP was usually the major form.
It is likely that this seasonal change of DP (or DRP) to PP proportions in runoff from pasture is strongly influenced by increased soil erosion in wetter periods caused by treading damage by livestock. It has been shown in many studies that grazing animals, particularly cattle, lead to a substantial increase in sediment loss in runoff from pasture (e.g., Sharpley and Syers, 1976; Lambert et al., 1985; Nguyen et al., 1998). Highest losses occur in winter and spring when soils are at their wettest, runoff volumes are highest, and treading damage at its most extreme (Lambert et al., 1985; Elliott et al., 2002). The extent of sediment loss caused by treading damage has also been shown to be influenced by the degree of slope (Nguyen et al., 1998), and the degree of slope has also been correlated with loss of P in runoff (Ahuja et al., 1982). In addition, several studies have shown that as erosion increases, the proportion of P exported in runoff in the form of PP increases as well (Sharpley et al., 1993, 1994). It seems likely, then, that the differences between the proportions of DP (or DRP) to PP in runoff from the hill-country and lowland catchment studies discussed above are to a fair degree due to differences in degree of slope, and the influence that this has on the amount of soil erosion, particularly that caused by stock treading. As the concentration of suspended solids in overland flow increases, DRP concentration tends to decrease due to adsorption onto the suspended material (Sharpley et al., 1981; McDowell et al., 2003a). Conversely, increased proportional losses as DRP in drier periods may be influenced by such factors as leaching of soluble P from plant material (Sharpley et al., 1992, 1995), lower uptake of soil solution P by plants (Rajan, 2002), and lysing of soil microorganisms (Turner and Haygarth, 2001).
All the above studies investigated losses of P from grassland sites. Hansen et al. (2000) investigated the loss of sediment and P, in snowmelt runoff from 66-m2 plots under different tillage systems, on a Clarion silt loam (Typic Hapludoll) soil, located in Scott County, Minnesota, USA. The three tillage systems compared were moldboard plow, chisel plow, and ridge till. Residue cover was 10, 40, and 93%, while surface roughness was 1.0, 0.76, and 0.12 cm for the moldboard plow, chisel plow, and ridge till treatments, respectively. In contrast to what might be expected from cultivated soil, DRP was the dominant form of P loss in snowmelt runoff, averaging 75% of TP loss. Average TP losses over a total of eight runoff events in two years were about 1.3, 1.1, and 0.4 kg ha–1 for the ridge till, chisel plow, and moldboard plow treatments, respectively. These differences were related to greater snowmelt runoff, the accumulation of P at the surface, and the leaching of P from the greater amounts of crop residue from the ridge till and chisel plow treatments. Snowmelt runoff does not cause substantial soil erosion because snow melts gradually, and because soil detachment is limited when soil is frozen (Hansen et al., 2000).
It would therefore appear that the answer to the question, Which form of P is predominant in surface runoff from agricultural land, dissolved or particulate?, is that it depends very much on the individual circumstances. It does appear that the presence of grazing animals may have a strong influence on the ratio of DP to PP in overland flow, but that this is modified particularly by the type of stock and topography (degree of slope). Nash and Halliwell (1999) and Sharpley et al. (2000) also emphasized the importance of timing of animal grazing in relation to runoff events. In pastoral systems, the main physical influence seems to be the treading damage caused by stock, which decreases infiltration of water into the soil, and exposes and loosens the soil surface. However, as indicated by the results of Hansen et al. (2000) above, it appears to be the exposure of soil to the erosive nature of rainfall impacting the soil surface and causing the suspension of soil particles that is one of the major driving factors for increasing the proportion of PP in surface runoff from grassland. This is in turn influenced by the topography, not only as hillslopes are more susceptible to treading damage, through hoof slippage for example, but also that rainfall intensity tends to increase with increasing altitude (i.e., in hillier country). The results of Heathwaite and Dils (2000) in turn indicate that the filtering of overland flow by pasture as it flows down a hillslope may also affect the proportion of DP to PP, but the degree to which this occurs will be a function of several factors, such as slope length, length and density of pasture, and volume and/or channeling of surface runoff (Smith, 1987; Muscutt et al., 1993; Daniels and Gilliam, 1996).
Clearly, there are some circumstances where one particular form of P is indeed predominant. For example, most studies on P loss from tilled soil show that most of the P exported is in the particulate fraction, usually attached to eroded soil particles (e.g., Schuman et al., 1973; Sharpley et al., 1993; Hodgkinson, 1996; Catt et al., 1998; Cox and Hendricks, 2000). However, as highlighted by this review, this is not always the case, and the literature has accumulated many conflicting generalizations, for example, "it is well established that the major pathway for P transport from agricultural soils is through soil particulate loss, particularly via runoff" (Ampontuah et al., 2001), "phosphorus losses from agriculture occur mainly due to the effect of soil erosion" (Strauss and Mentler, 1998), "phosphorus runoff from pasture is usually dominated by soluble forms" (Nash and Halliwell, 1999), and "surface runoff from grassland carries little sediment and is, therefore, generally dominated by dissolved P" (Sharpley et al., 2000). This review highlights the usually very site-specific nature of P losses in runoff, and that the fractionation of P in runoff from a given land use (e.g., "pasture"), can be diametrically opposite from one site to another. Such generalizations are thus shown to be of little value in increasing our understanding of P transport dynamics, and may even be counterproductive in some instances. A high degree of specific detail therefore needs to be given whenever P losses in runoff are being described.
On top of the factors mentioned above, the issue of forms and amounts of P in runoff from agricultural land is further complicated and confused by losses directly resulting from the addition of phosphate fertilizers or manures (Table 1). It is the view of the authors that the direct influence of fertilizer and manure application has been underemphasized, and in fact totally ignored in many otherwise in-depth studies of P runoff. This issue is examined in more detail in the next section.
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INFLUENCE OF RECENT FERTILIZER APPLICATION |
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TOPABSTRACTINTRODUCTIONFORMS OF PHOSPHORUS IN...INFLUENCE OF RECENT FERTILIZER...OVERALL COMMENTSDIFFERENCES BETWEEN TYPES OF...MECHANISMS FOR MANAGING EVENT...CONCLUSIONSREFERENCES |
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Much of the work done in this area has been conducted in New Zealand and Australia, probably reflecting the relatively high importance of agriculture to both the economy and water quality in these two countries. For the same reason, most published work deals with runoff from grazed pastures, rather than from cultivated land. Sharpley and Syers (1976) compared P losses in surface runoff from grazed and ungrazed plots on permanent pasture at Massey University, New Zealand, on Tokomaru silt loam soil, a Typic Fragiaqualf developed in siliceous loess, with or without the application of single superphosphate (SSP) at the rate of 50 kg P ha–1, over a 4-mo period. Overall losses of DIP (molybdate-reactive P, <0.45 µm) from the fertilized plots were approximately four to five times greater than those from the corresponding unfertilized plots. The proportion of added fertilizer lost in surface runoff from plots that were undrained ranged from 5.6 to 6.7%, of which between 48 and 52% was as TDP. For two plots that were mole-drained, 1.0% of fertilizer P was lost, of which 70% was as TDP. Fertilizer addition led to a significant increase in DIP and PP concentration in surface runoff, in the first week following fertilizer application. Mean DIP concentration in surface runoff increased to a peak of about 3.3 mg L–1 after 3 d, declining over the following 6 wk, though it remained higher than the background DIP concentration throughout the monitoring period. Mean PP concentration increased to a maximum value of 2.8 mg L–1 in the first week after fertilizer application, then decreased much more gradually than that of DIP from the same plot, and was still significantly higher than the background PP concentration almost 12 wk after the fertilizer was applied. This pattern of a sharp increase in P losses following soluble fertilizer applications, tailing off over a period usually of a few months, has been observed in every subsequent study.
Even greater increases in P concentration in runoff were measured at Massey University by Sharpley and Syers (1979a) following aerial application of SSP at 30 kg P ha–1 to a pasture catchment, and in further studies on the same runoff plots in a comparison of P runoff losses from solid and dissolved SSP at 50 kg P ha–1 (Sharpley and Syers, 1983). Amounts of total fertilizer P lost from all these studies ranged from up to about 3.8 to 11.5% of the amounts of the P applied, equivalent to exports of up to 3.25 to 7.1 kg P ha–1 yr–1. In a similar study, Sharpley and Syers (1979b) estimated that up to 4.8 and 8.8% of the P applied as SSP at 50 kg P ha–1 was transported in surface runoff as TDP and TP, respectively, or 2.4 and 4.4 kg P ha–1, over the course of a year. Some authors suggest caution should be exercised in extrapolating results from runoff plots to whole watersheds (Sharpley et al., 1978; Sharpley and Syers, 1979b), while others have found no significant difference when comparing P losses in simulated rainfall runoff from microplots compared with natural runoff from much larger paddock-scale plots (Cornish et al., 2002). Nevertheless, these data indicate that a substantial proportion of the P lost from pasture in surface runoff can be derived from recently applied fertilizer. In the context of typical background levels of P loss from pastoral land (e.g., around 1.3 kg P ha–1 yr–1 in New Zealand; Gillingham and Thorrold, 2000), these amounts are very large indeed.
McColl and Gibson (1979a)(1979b) measured P concentrations in runoff from unconfined collectors in two small (approximately 400 m2) hill-pasture enclosures grazed by sheep. Single superphosphate was applied in spring at about 51 kg P ha–1 to the top part of the enclosures. Fertilizer application caused a large increase in P concentrations in runoff from the enclosure that had recently been grazed (mean grass length 2.6 cm), but not from the ungrazed enclosure (mean grass length 10 cm). In the grazed enclosure, TP concentrations in runoff samples from the first post-fertilizer rain event were about 22 times higher than those in the runoff from the pre-fertilizer event, about 26 mg P L–1 compared with about 1.2 mg P L–1. The P concentrations decreased rapidly after the first post-fertilizer runoff event, 4 d after SSP application, but the fertilizer effects persisted until the third post-fertilizer event, which occurred 20 d after SSP application. This decline is somewhat faster than those typically measured in the studies of Sharpley and coworkers described above. However, those experiments were measuring P losses from plots to which fertilizer had been directly applied, whereas in the McColl and Gibson (1979b) experiment, fertilizer was deliberately not spread in the immediate vicinity of the collectors. The term "immediate vicinity" was not defined by these authors, but is assumed to mean within a meter or so. Nevertheless, the influence of fertilizer application was still very significant. The post-fertilizer runoff events accounted for less than 1% of the total annual downslope movement of water, but about 35% of the annual P movement, in the recently grazed pasture (McColl and Gibson, 1979c).
Gillingham et al. (1997) conducted an experiment comparing P runoff from microplots using simulated rainfall in the hill-country catchments at Whatawhata and at Waipawa in the Hawkes Bay region of New Zealand. At both sites, DRP in runoff 5 to 6 d following P fertilizer application (as SSP or monocalcium phosphate, MCP) was very high compared with that from the unfertilized plots. Mean DRP concentrations at Whatawhata increased from 0.06 to 6.4 mg L–1 while at Waipawa the mean concentrations increased from 0.46 to 31.5 mg L–1. The mean concentrations declined from these peaks quite rapidly, but the effects of the fertilizer applications on DRP in the runoff water were still apparent 90 to 110 d after the fertilizer was spread. This is much longer than that found by McColl and Gibson (1979a)(1979b). Obviously, local soil and other conditions will affect the length of time that recently applied fertilizer effects last; but from this review, a period of between around 30 to 50 d appears to be most typical for the most significant proportion of P losses. A notable point that arises from this observation is that there are relatively few areas in most temperate zones where at least one or two significant rainfall events would not occur during such a period, which would typically be in spring and/or autumn.
Other studies conducted outside of Australasia have found similar influences from recently applied fertilizer. Olness et al. (1980) compared nutrient losses from paired, gently sloping (3°) native grassland watersheds with Udertic Paleustoll and Udic Argiustoll silt loams soils, in Oklahoma, USA, under rotational or continuous grazing. Three days after ammonium phosphate fertilizer was broadcast at a rate of 75 kg P ha–1 on one of each paired watershed, an intense thunderstorm delivered about 94 mm of rain, generating between 3.6 to 5.4 cm of runoff from the watersheds. This led to an increase in TP in surface runoff of about 10- to 25-fold and in TDP of about 200- to 600-fold compared with losses from the unfertilized watersheds. Soluble P accounted for 79 and 65% of the TP exported in the first runoff event from the fertilized rotationally and continuously grazed watersheds, respectively. Runoff P concentrations decreased sharply with successive events, but after 2 mo and several runoff events the surface runoff P concentrations were still quite high, particularly the TDP fraction, being about 10 to 20 times greater than those from the unfertilized watersheds. Even 12 mo after fertilizer application, the runoff TDP concentrations were still 5 to 15 times greater than the long-term mean background concentrations. The authors estimated that over the course of a year, about 3 kg ha–1 of TP (79% TDP) and 5 kg ha–1 of TP (67% TDP) was lost in runoff from the fertilized rotationally and continuously grazed watersheds, respectively. The differences in P losses between the two grazing regimes over the course of the experiment were attributed to increased losses of PP associated with greater soil erosion under continuous grazing. As part of the experimental design, these watersheds were routinely overgrazed and generated on average about three times as much runoff as the rotationally grazed watersheds, and were severely eroded with active, steep-walled gullies.
A similar effect due to an intense rainstorm following shortly after fertilizer application was measured by Haygarth and Jarvis (1997), in an experiment conducted on 1-ha-sized grazed lysimeter plots on a clayey, noncalcareous pelostagnogley soil (Typic Haplaquept) in Devon, UK. Fertilizer (triple superphosphate, TSP) at the rate of 16 kg P ha–1 was added to some of the plots in spring. Over a period of 4 to 8 d following this application, more than 50 mm rainfall was recorded, which resulted in the highest P losses in runoff recorded during the experiment. Runoff in this experiment consisted of overland flow and subsurface flow to a depth of 30 cm. Mean DRP concentrations increased from 41 µg L–1 in a runoff event before fertilizer application to 252 µg L–1. A comparison of maximum values showed a 10-fold increase in DRP concentration from 131 to 1296 µg L–1. Comparisons of TP loading losses indicated even greater losses, with mean values of 92 compared with 4641 mg ha–1 h–1, with maximum values of 160 compared with 18510 mg ha–1 h–1. It was estimated, therefore, that for a storm that lasted about 30 h, losses may have been in excess of 0.5 kg P ha–1, 3% of the total P applied. The authors speculate that the high intensity rainfall may have caused erosion of PP, and thus contributed to the high losses of TP in relation to DRP. After excluding data from the runoff event immediately following the fertilizer application, TDP was found to contribute about 69% of TP in surface runoff and interflow to a 30-cm depth (Haygarth et al., 1998b).
Shuman (2002) measured DRP losses in simulated rainfall-runoff from 25-m2 turfgrass plots, on a Cecil sandy loam soil (Typic Kanhapludult) in Georgia, USA, with a 5° slope. Rainfall was applied and runoff collected 4, 24, 72, and 168 h after the application of monoammonium phosphate (MAP) fertilizer at 0, 5, and 11 kg P ha–1, in two successive years. Dissolved reactive P concentrations in the first runoff event were high, with an average over the two years of 0.75, 3.8, and 7 mg L–1 for the 0, 5, and 11 kg P ha–1 treatments, respectively. However, in this experiment the concentrations tailed off rapidly over the ensuing runoff events, and there was no statistical difference between treatments after 168 h, with average DRP concentrations being about 0.8 mg L–1. The large differences in length of time in the measurable effects of fertilizer applications on P loss in runoff in the experiments discussed above is another indication of the site-specific nature of P exports from agricultural land.
As noted previously, most studies on runoff from cultivated land indicate that most P losses in this situation are associated with the PP fraction. However, a few field studies have been conducted where fertilizer effects were measured against a control treatment, where this was not the case. Withers et al. (2001) compared P losses in surface runoff from 15 adjacent 32-m2 experimental plots with a 5° slope, on a silty clay loam soil (Argillic Albaqualf) developed over Old Red Sandstone in Herefordshire, UK, over a 2-yr period, with all plots being cropped to cereals each year. At the beginning of the experiment, TSP was surface-applied at a rate of 60 kg P ha–1. Three weeks following this, approximately 25 mm of rainfall fell on the site, producing approximately 16 L of surface runoff with an average DRP concentration of about 6.5 mg L–1, which constituted about 84% of the TP concentration. Subsequent (lower volume) runoff events in the month following fertilizer application produced much lower concentrations in runoff water. A cumulative load of about 82 mg P plot–1 was lost in surface runoff in the first three events, of which about 77% was TDP, compared with a total of 24 mg P plot–1 from the control treatments, of which only about 21% was TDP.
For the fourth monitoring period of the experiment the following year, treatments were applied to the seedbed surface, the preparation of which included rolling after being sown with winter barley (Hordeum vulgare L.). This had the effect of consolidating the soil surface, and meant that lower rainfall intensities were needed to generate surface runoff. In the first runoff event, which occurred within a few days of fertilizer application (TSP at 90 kg P ha–1), extremely high DRP concentrations were measured in runoff from the fertilizer-treated plots: 74 compared with 0.7 mg L–1 from the control plots. Treatment differences in P losses lasted for approximately 1 mo after application. Thus, even on conventionally cultivated soil, the presence of soluble P fertilizer on the soil surface can be the most influential determinant of the forms of P lost in surface runoff.
Zhang et al. (2003) compared DRP and TP losses from 30-m2 plots on two contrasting paddy soils under wheat (Triticum aestivum L.) in southeastern China, following application of inorganic P fertilizer (an N–P–K and SSP) at 0, 20, 80, and 160 kg P ha–1. One site (Anzhen) was on a loam clay soil with an impermeable clay layer down the soil profile, while the other site (Changshu) was on a silt loam soil with a relatively uniform and permeable profile. The time period between fertilizer application and the first runoff event was not recorded in the paper. However, significant differences were found both between sites and between rates of P application. Although the volume of runoff from the Anzhen site was approximately 25% greater than that from the Changshu site, P concentrations and loads were lower. In the control treatment, the majority of P exported was particulate (about 70–95% of TP). However, the proportion of P in runoff in the form of DRP increased with increasing rate of P application. At Anzhen, TP concentration in the first runoff event ranged from 1.13 mg L–1 in the control treatment to 4.13 mg L–1 in the 160 kg P ha–1 treatment, while DRP concentrations for these treatments were 0.26 and 2.76 mg L–1, respectively. The concentrations of TP gradually declined over the following runoff events, as did the proportion of TP in the form of DRP, in all treatments, which ranged from about 6 to 12% in the final, fifth runoff event, about 9 wk after the initial runoff event.
In contrast, concentrations and loads of runoff P from the fertilized plots were much greater at the Changshu site. While control plots values were similar to those at Anzhen, concentrations of TP and DRP in the first runoff event from the 160 kg P ha–1 treatment were 17.4 and 13.34 mg L–1, respectively. The concentrations remained higher than at the Anzhen site, even after six runoff events, and the proportion of DRP to TP remained relatively high throughout the experiment, in contrast to results from the other site. Overall, the proportion of TP exported from the fertilized plots at Anzhen in the form of DRP ranged from 9 to 34%, while that from the Changshu site ranged from 35 to 50%. These differences between the two sites were attributed to the difference in clay content and hence P adsorption between the two soils.
These results highlight the overriding influence that the application of soluble fertilizer can have on the characteristics of P runoff in the short term, and support the suggestion of Kleinman et al. (2002) that soil P contributes little to runoff P losses following recent surface application of labile P sources. They also support the suggestion that there is reduced contact between soil particles and P fertilizer granules under high rates of SSP application, allowing the fertilizer to remain a P source for a considerable time (Withers et al., 2003), although as demonstrated by Zhang et al. (2003), the extent of this effect will vary from site to site, depending in part at least on the soil characteristics, particularly clay content and degree of saturation of P sorption capacity.
Comparisons of Runoff from Inorganic Fertilizer and Manure Application
In many parts of the world, housing of stock over winter means that disposal of accumulated manures is a major problem, and as applications are usually made on the basis of N content, the amount of P that may be applied can be much greater than would be the case with an inorganic fertilizer application (Chadwick and Chen, 2002). Many studies have compared losses of P and other nutrients from inorganic fertilizers with those from organic manures and slurries. Nichols et al. (1994) measured losses of TDP and TP in simulated rainfall runoff from 9-m2 runoff plots under established tall fescue (Festuca arundinacea Schreb.), on a Captina silt loam (Typic Fragiudult) soil in Arkansas, USA, with a uniform 5° slope. Treatments were amendment with poultry litter or an N–P–K fertilizer at an equivalent rate giving 87 kg P ha–1, surface-applied, or with shallow incorporation to a depth of 2 to 3 cm by rotary tillage. Simulated rainfall was applied at 50 mm h–1, sufficient to produce 30 min of continuous runoff, 7 d after treatment application, and no significant difference was found in the volume of runoff coming from the various treatments. Total P concentrations in runoff were significantly lower from plots treated with poultry litter compared with inorganic fertilizer, with an average of 15.4 and 26.2 mg L–1, respectively. Approximately 67% of TP in the runoff from the poultry litter plots was in the form of TDP, whereas in runoff from the fertilizer treatment plots, >95% of TP was as TDP. The shallow incorporation method had no significant effect on the amounts of P lost from the fescue plots, due to inadequate turn-under of the surface-applied litter and fertilizer. As the depth of incorporation was very shallow, in an attempt to minimize damage to the fescue roots, it is probable that a significant part of the applied P that was incorporated into the soil was still at the effective depth of interaction for the overland flow (Sharpley, 1985), as well.
Heathwaite et al. (1998) compared N and P losses in runoff from sloping (15–20°), ungrazed grassland plots on weakly structured brown Cambisol soil of the Denbigh Association, receiving inorganic fertilizer, farmyard manure (FYM), and liquid cattle slurry. A compound N–P–K fertilizer supplying P as DAP was applied to some of the plots at the rate of 100 kg P ha–1. Simulated rainfall was applied to the plots at a rate of 22 mm h–1 for 35 min on four consecutive days, starting 3 d after the day of fertilizer application. Much greater TP concentrations were measured from the fertilizer plots (mean 15.3 mg P L–1) compared with those from the FYM and slurry treatments (1.76 and 0.82 mg P L–1, respectively). Total P export in surface runoff from the base of the 20-m hillslope plots receiving fertilizer averaged about 3.8 kg P ha–1, of which about 67% was in the form of DRP, compared with about 0.1 kg P ha–1 from the untreated control plots.
It could be said that applying artificial rainfall so shortly after fertilizer application is not a very realistic scenario, but in many if not most temperate parts of the world, rain events are unpredictable over any longer than a few days ahead, and so the coincidence of rainfall and fertilizer application is bound to happen from time to time. Preedy et al. (2001) conducted an experiment specifically designed to measure event-specific transfer of P from 30-m2 plots treated with inorganic and organic amendments, on the same site used by Haygarth and Jarvis (1997), over a 7-d period. Treatments were TSP, dairy slurry, and TSP + slurry, applied at the rate of 29 kg P ha–1, to grassland that was already wet, with further rainfall forecast. Water discharge collected from the plots came from overland flow and interflow to a depth of 27 cm, and accounted for 98% of the rainfall that fell on the plots over the 7 d. Concentration in runoff from the TSP and TSP + slurry treatments peaked at around 11 mg TP L–1 (67–68% as DRP), while those from the slurry treatment peaked at about 7 mg TP L–1 (20% as DRP), when rainfall was at its most intense (approximately 3 mm h–1) about 28 to 32 h into the experiment. This 4-h period accounted for 18% of rainfall and discharge, but produced 46% of the TP load from the TSP treatment, compared with 38% in the TSP + slurry, 33% in the slurry, and 27% in the control, respectively. It was estimated that over the course of the experiment, P loads from the TSP and slurry-treated plots were equivalent to losses of about 1.8 kg TP ha–1, while those from the TSP + slurry treatment were equivalent to about 2.3 kg TP ha–1 (approximately 6.2 and 7.9% of applied P, respectively), compared with only about 0.06 kg TP ha–1 from the control plots.
Gaudreau et al. (2002) measured nutrient concentrations in runoff from 6-m2 turfgrass plots, on a Boonville fine sandy loam (Chromic Vertic Albaqualf) soil in Texas, USA, treated with composted dairy manure applied at 50 and 100 kg P ha–1, and an unidentified inorganic P fertilizer applied at 25 and 50 kg P ha–1. Generally, differences in the TDP concentration between the control and the other treatments were relatively small, the latter being roughly one to three times greater. However, one runoff event happened to occur 3 d after treatment application, compared with a lag of 27 to 87 d for the other events. The control TDP concentration was 1.7 mg L–1, while those from the 50 and 100 kg P ha–1 manure treatments were 5.5 and 9.8 mg L–1, respectively, and those from the 25 and 50 kg P ha–1 fertilizer treatments were much greater at 16.6 and 30 mg L–1, respectively. Other field experiments reviewed in this paper have also demonstrated the coincidence of natural rainfall with fertilizer application (e.g., Olness et al., 1980; Nash et al., 2000b).
Given the expensive nature of field experiments and the unpredictability of natural rainfall, the use of rainfall simulators and laboratory microcosms is becoming more common in P transport research. Kleinman et al. (2002) compared losses of P in simulated rainfall-runoff from three different manures and DAP fertilizer applied at 100 kg P ha–1 to three soils packed into 0.2-m2 runoff boxes. The soils were a Buchanan (Aquic Fragiudult)–Hartledon (Typic Hapludult) association, Hagerstown (Typic Hapludalf), and Lewbeach (Typic Fragiudept), in two sets, one with low (12–26 mg kg–1 Mehlich-3 P) and the other with high background P levels (396–415 mg kg–1 Mehlich-3 P). All P amendments, with the exception of dairy manure on the Buchanan–Hartledon and Hagerstown soils, led to very significantly increased losses of P in runoff generated 72 h after treatment application. Dissolved reactive P in runoff water accounted for 64% of TP overall from the amended soils, compared with only 9% from the control soils. Losses of P were greatest from the DAP treatments, although they were generally not statistically different from those of the other amendments (p = 0.05). Losses of P were mostly, though in the case of dairy manure amendment, not always, greater from the high P soils compared with the low P soils. The differences between these two sets was most pronounced in the DAP treatments. If the average amounts of P lost from the DAP-treated soil boxes are scaled up directly, then the equivalent of 1.1 and 1.4 kg P ha–1 was lost as DRP and TP from the low P soils, and the equivalent of 4.1 and 6 kg P ha–1 as DRP and TP from the high P soils, respectively. Differences between the amounts of P lost from plots treated with inorganic P fertilizers compared with organic manures and slurries appear to be correlated with the relative amount of soluble P present in these P sources.
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TOPABSTRACTINTRODUCTIONFORMS OF PHOSPHORUS IN...INFLUENCE OF RECENT FERTILIZER...OVERALL COMMENTSDIFFERENCES BETWEEN TYPES OF...MECHANISMS FOR MANAGING EVENT...CONCLUSIONSREFERENCES |
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Surprisingly, despite the demonstrable significance of individual applications of fertilizer on P runoff from agricultural land, there have been very few studies comparing runoff from different fertilizer forms of P. We believe that this reflects a general lack of knowledge and/or interest in the wide range of solubilities, physical characteristics, and soil chemical reactions affecting different phosphate fertilizers. This topic is discussed in the following section.
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TOPABSTRACTINTRODUCTIONFORMS OF PHOSPHORUS IN...INFLUENCE OF RECENT FERTILIZER...OVERALL COMMENTSDIFFERENCES BETWEEN TYPES OF...MECHANISMS FOR MANAGING EVENT...CONCLUSIONSREFERENCES |
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Other studies have also shown massive losses of P from irrigation pasture, for example, up to 12.9 kg ha–1 yr–1 by Austin et al. (1996), up to 11.8 kg ha–1 yr–1 by Barlow et al. (2000), and up to 18.6 kg ha–1 by White et al. (2003). Extremely high losses tend to occur when soluble fertilizer is applied just before flood irrigation. Even when there was a delay of 34 d between application of monoammonium phosphate (MAP) fertilizer (at 40 kg P ha–1) and flood irrigation, White et al. (2003) measured a total load of 10.9 kg total reactive P ha–1 in runoff, which represented 25.6% of the applied P, compared with 0.6 kg total reactive P ha–1 from the unfertilized control treatment. In contrast, Bush and Austin (2001) suggested that a delay of 3 d or greater between applying SSP and flood irrigation would be sufficient to halve the concentration of P in runoff. However, their experiment involved sequential flood irrigations rather than a single irrigation with differing lags between fertilizer and irrigation application, as done by White et al. (2003), so it is likely that their results were affected by most of the soluble P present being exported in the first irrigation runoff. It is also possible that these two different P fertilizers behave differently under these conditions in terms of P release (Nash and Halliwell, 1999; Nash et al., 2003). These very contrasting results suggest that a suitable delay between spreading soluble P fertilizer and applying flood irrigation probably needs to be determined for different sites and fertilizers on an individual basis, at least until further research has been done.
Comparisons of Fertilizers of Varying Solubility
Sharpley et al. (1978) compared losses in surface runoff of DIP and PP derived from SSP and dicalcium phosphate (DCP) fertilizers applied at 50 kg P ha–1 to permanent pasture, on Tokomaru silt loam soil. After the fertilizers were applied, there was an immediate and substantial increase in the mean concentration of DIP in surface runoff (Fig. 1)
. The maximum DIP concentration of surface runoff from SSP application was 3.45 mg L–1 in the first runoff event 8 d after fertilizer application, a very similar result to the previous study (Sharpley and Syers, 1976), whereas that from application of DCP was 2.83 mg L–1, and the mean DIP concentration from the DCP-fertilized plot was always lower than that from the SSP-fertilized plots (Fig. 1). In contrast, the mean concentrations of PP in surface runoff after the DCP application were appreciably greater than those after SSP application (Fig. 2)
. The mean PP concentrations peaked at 4.65 and 2.83 mg L–1 in the runoff event immediately following the DCP and SSP applications, respectively.
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Bush and Austin (1999) compared P losses in runoff from irrigated grazed pasture bays on Goulburn loam soil (Sodic, Hypocalicic Brown Chromosol), at Tatura, VIC, Australia, treated with either SSP or Pivot Prolong (a granulated mixture of SSP and North Carolina phosphate rock, with a soluble P content of approximately 55%), at a rate of 44 kg P ha–1, and from untreated control bays. Phosphorus concentrations in runoff from the irrigation bays were measured over approximately two years, during which time the fertilizer treatments were applied three times. Unfortunately, there were large variations in runoff volumes within and between treatments (in some irrigations, insufficient water was applied to produce any runoff). Furthermore, not all data were adequately reported. Despite these shortcomings, examination of the overall mean total P loads removed from the bays over the 2-yr experimental period showed that losses of TP were equivalent to 16.3 kg P ha–1 for the control, 35.5 for the Prolong, and 53.4 for the SSP-treated bays, respectively.
These amounts are approximately in line with the relative water solubility of the TP content of the two fertilizers, and represent truly massive losses of fertilizer P to the environment. Even the untreated control bays in this experiment exported greater amounts of P per annum (average 8.1 kg TP ha–1) than most of the fertilized treatments in other authors' work reviewed in this paper, possibly caused by high losses of excretal P from recent grazing events.
Use of Slow-Release Forms of Phosphorus
In recent years, use of slow-release fertilizers such as direct-application phosphate rock (DAPR) has been proposed as a method to reduce loss of P in runoff (Sharpley and Withers, 1994; Nelson et al., 1996), but there has been relatively little experimentation done to test this outside of New Zealand, and more recently Northern Ireland and Victoria, Australia. Nguyen et al. (1999) compared P runoff from hill-country 0.5-m2 microplots at Whatawhata, New Zealand, treated with fully soluble fertilizers, SSP and TSP, slow-release DAPRs from Tunisia (Gafsa) and Egypt (Kosseir), and a 50:50 blend of the two DAPRs and TSP in a 60:40 ratio of TP, applied at a rate of 35 kg P ha–1. Surface runoff samples were collected from five simulated rainfall events over a 232-d period, and analyzed for DRP and PP. The results showed highly significant differences in the P concentration of runoff water and in the total amounts of P losses between the different treatments (Table 2).
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Much less P was transported in runoff from the plots as PP than as DRP, and the differences between treatments were also much less pronounced (Table 2). For example, amounts of PP measured in runoff on Day 3 after SSP or TSP application were 3.2- and 6.2-fold greater than those from the control, respectively. In contrast, amounts of DRP measured at Day 3 were 180 and 262 times higher from the respective fertilizer treatments than from the control. Amounts of PP in runoff from the DAPR treatments ranged from 0.9 to 1.9 times that from the control, over the course of the experiment. Overall, 1.73 to 2.32% of the P applied as SSP or TSP was lost in surface runoff as DRP and PP combined, compared with 0.07% of the P applied as DAPR. The total amount lost from the DAPR–TSP treatment amounted to about 0.8% of the P applied. The loss from the SSP and TSP treatments equates to a loss of about 0.6 to 0.8 kg P ha–1, compared with 0.02 kg P ha–1 from the DAPR treatments. A similarly designed field trial, conducted in Northern Ireland and using the same type of rainfall simulator, yielded similar trends. Concentrations of DRP and TP in a runoff event a few days after application of DAP were 0.037 and 0.042 mg L–1, respectively, compared with only 0.002 and 0.004 mg L–1 with DAPR application, and 0.0015 and 0.0035 mg L–1 for the no-P control (Quin et al., 2003). Over the long term, such differences in P loss from an agricultural system would represent a significant financial saving to a farmer, as well as greatly reducing the potential for environmental pollution. The implications for use of DAPRs on flood-irrigated perennial pastures in light of the very high losses discussed above are obvious, although seemingly this is an area that has yet to be investigated.
A follow-up study to the New Zealand field experiment was conducted on sieved topsoil taken from trial sites with a long-term history (>20 yr) of either SSP or DAPR application, and repacked into microplots (1 x 0.5 x 0.1 m) in the laboratory (Nguyen et al., 2002). Fresh applications of SSP or Gafsa DAPR at a rate of 35 kg P ha–1 were made to the microplots, and surface runoff induced using rainfall simulators 3, 10, and 32 d following fertilizer application. As with the field study, the most DRP was lost in the initial runoff event from the SSP-treated plots, ranging from 0.22 to 0.25 kg P ha–1 equivalent. Runoff amounts of DRP from the SSP treatments were still 22 to 30 times greater than those from the untreated controls after 32 d however, in contrast to the field study where the equivalent loss was about nine times greater than the control (Nguyen et al., 1999). Significantly greater amounts of PP were lost in this experiment compared with the field trial, presumably due to the lack of plant cover (Table 1). The fertilizer history of the soil had no significant effect on the amount of DRP or PP lost from the fertilizer applications, with 1.13 to 1.26% of the SSP applied being removed in runoff over the 32-d period of the experiment. Overall, this laboratory experiment confirmed the much lower amounts of P export in runoff from plots treated with Gafsa DAPR compared with highly soluble SSP fertilizer that was found in the earlier field experiment. It also indicates that previous applications of different P fertilizers do not influence the dynamics of P loss in runoff, once the fertilizer P has reacted with soil constituents, and becomes effectively indistinguishable from "native" soil P, further supporting the assertion of Kleinman et al. (2002), mentioned above.
A recently published similar study, comparing DAPR (in this case Sechura) with SSP, applied at 23 kg P ha–1 equivalent to soil box microcosms angled at 5°, using Lismore stony silt loam (Udic Ustochrept) soil, also found large differences in losses of P in overland flow from simulated rainfall applied at 15 mm h–1 (McDowell et al., 2003b). Over the course of the 184-d experiment, the flow-weighted mean DRP concentration in overland flow was 0.743 and 0.106 mg L–1 for the SSP and DAPR treatments, respectively, while that of the PP fraction was 0.239 and 0.122 mg L–1, respectively. However, in the first runoff event 7 d after fertilizer application, the DRP concentration in runoff from the SSP treatment was 2.78 compared with 0.16 mg L–1 from the DAPR treatment, while that of PP was 0.41 and 0.22 mg L–1, respectively. The concentrations of TP in the first runoff event (mg L–1) were 3.57 for the SSP, 0.39 for the DAPR, and 0.27 for the control (0 P) treatments, respectively. The concentration of P in overland flow from the SSP treatments was significantly greater than that of the DAPR and control treatments for approximately 60 d following fertilizer application. A significant point from this experiment is that these large differences in P exports were found under a rainfall application rate much lower, and more commonly occurring in nature, than that used in most other work using rainfall simulators, which is more often around 50 to 75 mm h–1 (e.g., Sharpley et al., 1981; Pote et al., 1999; Cox and Hendricks, 2000).
While noting again that caution ought to be taken when extrapolating losses from small-scale experiments to larger paddock and farm scales, the results of these experiments comparing DAPR with soluble P fertilizers are nevertheless of considerable importance. As noted above, under flood irrigation and extreme rainfall-runoff events, amounts of P that are not just environmentally but also economically important can be lost in runoff following recent applications of soluble P fertilizer. Under less extreme conditions, it is considered that usually less than about 5% of the P applied in fertilizer is lost in runoff (Sharpley et al., 1995). However, a loss of a few percent may still be sufficient to raise surface water P concentrations to undesirable levels (e.g., McColl et al., 1975; Nelson et al., 1996; Wilcock et al., 1999). A significant proportion of the increasing amounts of P export occurring from agricultural land to surface water is attributed to increases in soil P levels, through an imbalance between P inputs and outputs in modern farming systems (Sharpley et al., 1999; McDowell and Sharpley, 2001). However, r
