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Dep. of Geography, Royal Holloway, Univ. of London, Egham, Surrey, UK TW20 0EX
phil.haygarth{at}bbsrc.ac.uk
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ABSTRACT |
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Abbreviations: RP, reactive phosphorus • RP (<0.45), reactive phosphorus <0.45 µm • RP (>0.45), reactive phosphorus >0.45 µm • TP, total phosphorus • TP (<0.45), total phosphorus <0.45 µm • TP (>0.45), total phosphorus >0.45 µm • UP, unreactive phosphorus • UP (<0.45), unreactive phosphorus <0.45 µm • UP (>0.45), unreactive phosphorus >0.45 µm
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMethodsResultsDiscussionConclusionsREFERENCES |
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The transfer of P from agricultural land can occur through surface or subsurface pathways, although the capacity of most subsoils to fix inorganic P has meant that subsurface transfer has traditionally been perceived to be of minor importance (Baker et al., 1975; Burwell et al., 1977; Sharpley and Syers, 1979). However, it is now recognized that P can be exported through subsurface pathways at levels that can cause problems for water quality (Foy and Dils, 1998). This phenomenon is not restricted to waterlogged soils (Khalid et al., 1977) or sandy-textured soils under heavy fertilization as traditionally thought (Ozanne et al., 1961; Breeuwsma and Silva, 1992), but includes many soil types, especially clay soils that are susceptible to cracking and preferential flow (Simard et al., 1998; Stamm et al., 1998). However, despite the accepted role of subsurface pathways in P transfer, most evidence is derived from studies on tile drainage at the plot or field scale (e.g., Sawhney, 1978; Turtola and Jaakkola, 1995; Grant et al., 1996; Haygarth et al., 1998; Simard et al., 1998; Stamm et al., 1998). This approach may not give a true reflection of subsurface P transfer under natural conditions, because artificial drainage creates preferential flow pathways that minimize contact with the subsoil and can strongly alter the processes controlling P release to drainage water. In addition, these studies only represent heavy soils that require artificial drainage and do not include lighter, more freely draining, soils.
Thus, there is little direct field information on the forms and concentrations of P in water draining through soil that is not artificially drained. This information is essential, to understand the impacts of subsurface drainage on water quality and to allow the development of strategies for the control of agricultural P pollution. The aims of this study were to determine P forms and concentrations in leachate water at the soil profile scale under field conditions in the absence of artificial drainage. To achieve this, we used large-scale monolith lysimeters to monitor P in leachate from four grassland soil types.
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Methods |
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TOPABSTRACTINTRODUCTIONMethodsResultsDiscussionConclusionsREFERENCES |
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Analytical
The leachate was analyzed for a range of operationally defined P fractions. Total P (TP) was determined by a sulfuric acid–persulfate digest adapted from Eisenreich et al. (1975) and described by Rowland and Haygarth (1997), except that we used 0.15 (± 0.01) g of potassium persulfate reagent, rather than the incorrect amount stated in the published method. Reactive P (RP), generally considered to be inorganic orthophosphate, was determined using a Tecator 5020 flow injection analyzer with an autosampler (Method Application ASN 60-03/83, Tecator Ltd, Höganäs, Sweden) using a molybdenum blue reaction at 690 nm (Haygarth et al., 1998). The difference between TP and RP is unreactive P (UP).
The analyses were conducted on (i) unfiltered samples and (ii) samples filtered through a 0.45 µm cellulose-nitrate membrane filter. The qualifier <0.45 is used to describe P forms in filtered samples (traditionally termed dissolved P), while >0.45 is used to describe the difference in P between total and filtered P fractions (traditionally termed particulate P). Fractions with no qualifier describe P forms in unfiltered samples (Haygarth and Sharpley, 2000).
Statistical
Differences in leachate TP between soil types during the 1994 through 1995 drainage year were analyzed statistically using one-way analysis of variance tests. Tests were conducted on (i) TP concentrations on individual sampling events (ii) mean TP concentrations over the entire drainage year (1994–1995) and (iii) TP loads over the entire drainage year. Significant differences between individual soil types were determined by comparing the least squared difference of means value (LSD), obtained from the original ANOVA test, with the difference in mean values between any two soil types. If the LSD value was less than the difference in means for any two soil types, this indicated that mean values for the two soil types were statistically significantly different.
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Results |
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TOPABSTRACTINTRODUCTIONMethodsResultsDiscussionConclusionsREFERENCES |
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Loads and Concentrations of Phosphorus in Leachate
Temporal Trend in Reactive Phosphorus
Reactive P (<0.45) concentrations were determined across two drainage years. Mean concentrations across the full sampling period between September 1993 and June 1995 ranged between 53 µg L-1 for the sandy loam to 121 µg L-1 for the silty clay (Fig. 1)
. For each soil type, the mean concentrations were larger than the flow-weighted mean concentrations of RP (<0.45), implying that the mean data were distorted by a few high concentration events. Concentrations of RP (<0.45) for individual sampling events frequently exceeded 100 µg L-1, especially for the silty clay, which represents the single most common soil type in the UK (Haygarth et al., 1998). Concentrations were generally consistent throughout the drainage year, although elevated concentrations occurred in the late spring periods (April–May). At these times, maximum concentrations of >200 µg L-1 were recorded for all soil types, indicating that particular processes were affecting P in leachate irrespective of soil type (Fig. 1). The clay-textured soils had greater RP (<0.45) concentrations in leachates than the sandy-textured soils. The sandy loam leachates consistently had the smallest concentrations of RP (<0.45). We suggest that this effect might be partly related to the soil pH levels, which were higher in the top 30 cm of the sandy-textured soils (Table 1).
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Total Phosphorus Concentrations
Total P concentrations in leachate were determined during the 1994 through 1995 drainage year. Leachate collected during this drainage year was analyzed less intensively, but for the full range of P fractions. The less intensive sampling resulted in the mean concentrations for the sand and the clay loam being skewed by a single high concentration event (Fig. 2)
. As a result, flow-weighted means provide a better indication of the concentrations of P fractions for these soils during this drainage year. A summary of the mean concentrations of each fraction is shown by soil type in Table 2
. Total P concentrations during 1994 through 1995 showed a similar temporal and between-soils trend to those for RP (<0.45) for the previous year (Table 2, Fig. 2). Flow-weighted mean concentrations of TP in leachate ranged between 53 µg L-1 for the sandy loam, to 163 µg L-1 for the silty clay.
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Total Phosphorus Loads
The load of TP exported from the lysimeters was calculated by multiplying the TP concentration at the time of sampling with the volume of leachate preceding the sample. This gave a very imprecise estimate of the TP loads, because of the infrequent sampling times. However, these values are useful to give an idea of the magnitude of P export and to determine the proportional contribution of the various P fractions to the TP load. The load values for each fraction are presented in Table 3
. During the 1994 through 1995 drainage year, the mean TP export was <0.5 kg ha-1 yr-1 for all soil types. Total P exports for individual lysimeters ranged between 202 and 594 g ha-1. These exports typically represented <1% of the annual fertilizer applied (40 kg ha-1) and a minute fraction of the total reservoir of soil P. Total P loads were remarkably similar for all soil types except the sandy loam, which was about 60% of the others (Table 3). However, statistical analysis of TP loads for the 1994 and 1995 drainage year showed that differences among soil types were not significant (P > 0.05), even for the sandy loam soil. This was a consequence of the variability among lysimeters of the same soil type. Total P export from the silty clay was similar to the clay loam and the sand, even though drainage through the silty clay soil was considerably less.
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Most of the RP occurred in the <0.45 size range (38–63% of the TP) (Tables 3 and 4). Thus, RP (<0.45) dominated the P export from all soil types. The clay loam leached the greatest proportion of RP (<0.45), while the sandy loam leached the smallest proportion. Reactive P (>0.45) was mostly small and only accounted for a significant proportion of the TP exported from the sandy loam (26%).
In contrast to RP, which occurred mostly in the <0.45 fraction, UP was present mainly in the >0.45 fraction, accounting for up to one-third of the TP export. Indeed, UP (>0.45) dominated the TP (>0.45) fraction from all soil types except the sandy loam (e.g., for the silty clay, as shown in Fig. 3) . In particular, UP (>0.45) exported from the silty clay and the sandy soil represented about 75% of the TP (>0.45) export. Concentrations of UP (<0.45) were generally small from all soil types (flow-weighted mean concentrations between 5 and 14 µg L-1) and only accounted for 6 to 17% of the TP exported, although this represented more than 50% of the UP export from the clay loam (Table 4). However, a clear temporal trend in UP (<0.45) was apparent, with maximum concentrations of up to 71 µg L-1 recorded from all soil types in the springtime (April–May) (Fig. 4) . Therefore, UP (<0.45) appears to have a pronounced seasonal cyclicity.
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Discussion |
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TOPABSTRACTINTRODUCTIONMethodsResultsDiscussionConclusionsREFERENCES |
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Concentrations of RP (<0.45) did not fluctuate greatly about the mean value. This appears to indicate that sorption processes between the solid soil and the leachate water were acting to maintain an equilibrium P concentration in the soil water. The exception to this was during the late spring period (April–May), when maximum P concentrations of >200 µg L-1 were detected, which were similar across all soil types. The reasons for the peaks in RP (<0.45) concentrations at these times are uncertain, but may represent a range of processes, including preferential flow, incidental losses of surface-applied fertilizer, or biological P release through wetting and drying cycles. For example, preferential flow during the spring would result in the rapid movement of high P water from the surface down through the soil profile and, therefore, a breakdown of the chemical equilibrium processes operating during the winter drainage period. In addition, application of fertilizer to the soil surface during the springtime may increase the potential source of P available to drainage water. A combination of these two factors may have resulted in the high P concentrations determined in the spring period.
The sandy loam had the smallest P export. This may be related to the presence of a chalk layer at depth, as calcium in the soil is known to strongly fix inorganic P by precipitation (Frossard et al., 1995). Calcium in solution appeared to cause the precipitation of inorganic P in the drainage outlets and collection vessels, which was observed as a white insoluble precipitate. Although this was not analyzed, it is likely that these were carbonates, given that the leachate had passed through calcium carbonate in the soil profile. This reduction of inorganic P in solution may have resulted in an overestimation of the importance of other P forms in drainage water from the sandy loam soil and must be considered when interpreting the data from this soil.
In addition, problems were encountered with the silty clay soil. A clay layer at 50-cm depth caused saturation of the topsoil under high rainfall, impeded vertical drainage and caused overspilling of the lysimeters: this would not happen in a true field situation. This resulted in mean drainage from the four replicate lysimeters to be about 50% of that from the other three soil types. Therefore, the data obtained from this soil are of arguable validity.
Despite wide variations in the extractable-P contents of the remaining soil types, there were remarkably small differences in the TP exported in leachate (about 0.4 kg ha-1 yr-1) (Table 3). These differences were not statistically significant, because of the large variations in P concentrations within the four replicates of each soil type. Mean annual export of RP (<0.45) was about 0.3 kg ha-1 from all soil types. Similar exports have been reported by Turtola and Jaakkola (1995) for heavy clay soils under barley and grass (0.32 and 0.34 kg ha-1 yr-1, respectively), although much lower losses were reported by Burwell et al. (1977) and Sharpley and Syers (1979). Garwood and Tyson (1973) recorded P export of 0.3 to 0.4 kg ha-1 yr-1 from lysimeters containing the same sandy loam over chalk soil used in this work. In addition, they reported P concentrations ranging between 0.1 and 0.7 mg L-1, but stated that these losses were insignificant, reflecting the traditional agronomic view that P leaching was unimportant in context with the agronomic system.
There were no relationships between total P, Olsen-extractable P, water-extractable P, or CaCl2–extractable P in the soil and P fractions in drainage water. Indeed, lysimeter soils that might be classed as P deficient on the basis of simple soil P tests, actually leached more P than soils with much greater P contents. An initial interpretation might suggest that this is in contrast to the findings of Heckrath et al. (1995), who discovered that for arable soils with a history of fertilizer application, the Olsen-P status was related to the inorganic P concentration of drainage water, with an accelerated change point at 60 mg Olsen-P kg-1. However, there are two important differences between the study of Heckrath et al. (1995) and the current study: (i) Heckrath et al. compared P contents in one soil type whereas our range of Olsen-P values are represented by different soils; and (ii) the current systems are grasslands which do not represent the high range of Olsen-P concentrations that are common in arable soils.
Therefore in grassland soils, other factors may be proportionately more important in determining leachate P concentrations. In particular, the hydrology of the soils may exert a strong control on P transfer through subsurface pathways. At a simple level, the amount of rainfall will determine the P export from a soil of a given P status. This was demonstrated by Baker et al. (1975), who found that P export varied considerably from year to year depending on the rainfall and, therefore, the amount of runoff. In addition, the response of the soil to rainfall can determine the P transferred, especially through the extent of preferential flow through the soil (Simard et al., 1998). An alternative explanation of the lack of correlation with soil P pools may be the controlling influence of soil pH. The clay-textured soils (pH 5.7 and 6.5) leached greater concentrations of P than the sandy-textured soils (pH 7.0 and 7.3), suggesting that P leaching may be restricted at higher pH levels. This may be linked to the relationship with calcium discussed previously.
Forms of Phosphorus
Phosphorus export from all soils was dominated by reactive P in the <0.45 µm size fraction, as found in other studies (Sharpley and Syers, 1979; Heckrath et al., 1995; Chapman et al., 1997). This fraction is often assumed to represent true dissolved orthophosphate, although there is a wealth of evidence that this is not necessarily the case. Firstly, the acidic nature of the Mo-blue color reaction for the determination of RP can result in the hydrolysis of labile sugar phosphates, although more recent evidence suggests that this is negligible if samples are rapidly analyzed (Denison et al., 1998). Secondly, it is now clear that the <0.45 µm fraction does not represent dissolved P, but actually contains a continuum of particle sizes containing P bound to inorganic and organic colloids, which contribute to RP (e.g., Haygarth et al., 1997; Sinaj et al., 1998). This has implications for the transport of P through the soil, because many UP and colloidal P forms are less strongly sorbed in the soil than inorganic P, which is readily fixed and prevented from leaching by precipitation with Ca, Fe, and Al, or sorption to clays and other soil particles (Frossard et al., 1995). Therefore, these forms can move easily through the soil and escape to waters (Rolston et al., 1975; Frossard et al., 1989; Chardon et al., 1997).
A substantial proportion of the P export from all soil types was in >0.45 µm forms. The potential for P (>0.45) transport through macropores was demonstrated by Simard et al. (1998), while Dils and Heathwaite (1996) detected P concentrations in macropore flow through agricultural grassland soils of >1 mg L-1, which was dominated by TP (>0.45). In addition, substantial subsurface concentrations of TP (>0.45) have been recorded indirectly in drain outflow (Turtola and Jaakkola, 1995; Grant et al., 1996; Haygarth et al., 1998). These were found to contribute up to about 70% of the TP exported in artificial drainage from loamy arable soils (Grant et al., 1996) and up to 0.2 kg ha-1 yr-1 in from grassland plots (Turtola and Jaakkola, 1995). The large amounts of TP (>0.45) present in subsurface drainage water indicate that erosion mechanisms within macropores may contribute to P transfer. The likely mechanism of release is simple physical detachment of particles from the walls of macropores and other preferential drainage pathways. The importance of TP (>0.45) transport confirms that process terms such as P leaching (which is synonymous with all subsurface P transport, but actually means the release and movement of exclusively dissolved P), can be very misleading (Haygarth and Sharpley, 2000).
The extent to which preferential flow through the subsoil is likely to determine the impact of subsurface P transfer on catchment water quality is open to question, but is likely to depend on the degree of hydrological connectivity with the receiving surface drainage channels. Evidence is accumulating that excellent connectivity exists between subsoils and receiving streams, especially for undisturbed, permanent grassland, through well-defined macropore and soil pipe systems. For example, Goulding and Webster (1992) noted the importance of preferential flow through soil pipes in experimental field soils, while Nieber and Warner (1991) stated that in many instances, soil pipes contribute most of the total subsurface stormflow from an idealized hillslope. Preferential flow paths may be responsible for the majority of the subsurface P transfer. For example, Jensen et al. (1998) discovered that orthophosphate transfer only occurred through wide-aperture macropores in structured soils, despite water flow not being restricted to the same macropores. However, this may be due to inorganic P adsorption to the pore walls, and might not necessarily be the case for particulate and unreactive (non-orthophosphate) P forms that are less susceptible to sorption. The large amounts of these non-orthophosphate forms determined in subsurface drainage in this study indicates that much of the P is not restricted to large diameter flow pathways and has the potential to move through a range of water pathways.
A large proportion of the TP export was unreactive and, therefore, likely to be mostly in an organic form. Soil scientists have traditionally neglected organic P as an algal available P source, because orthophosphate was considered to be the main bioavailable form. However, UP forms are potentially available to algae, especially in the <0.45-µm size range, after hydrolysis and release of inorganic P by the action of phosphatase enzymes (Jansson et al., 1988; Turner and Haygarth, 2000). Information on the transfer of UP forms is limited, because they are very difficult to separate and detect, without the availability of time consuming and expensive analytical techniques (Espinosa et al., 1999). However, organic P forms vary considerably in their bioavailability and behavior in the soil. For example, nucleic acids and nucleotides appear to be rapidly broken down, whereas inositol-bound P forms are much more recalcitrant and unavailable for biological uptake (Turner and Haygarth, 2000). As a result of the variable bioavailability of organic P, it is important to have some knowledge about the specific forms present in drainage waters, in order to gauge the impact of P transfer on eutrophication. Although it is currently difficult to quantify the bioavailability of organic P to algae, new techniques using enzymes to characterize specific organic P forms appear to offer the best prospect of this becoming routine (Turner and Haygarth, 2000). Until this is possible, the entire organic P fraction must be considered when assessing the impact of P transfer from the land.
Unreactive P (>0.45) accounted for the majority of the TP (>0.45) export. Information on the specific forms of P constituting the UP (>0.45) fraction is scarce, but probably represents P held within soil particles, organic P bound to soil particles, and bacterial cell debris (Hannapel et al., 1964). For example, Heathwaite et al. (1990) stated that about 80% of the TP export from a small mixed agricultural catchment in southwest England was in the form of organic P bound to soil particles. Stevens and Stewart (1982) found that alkali-soluble TP (>0.45) isolated from river water draining into Lough Neagh, Northern Ireland, resembled humic acid, with P associated with iron and organic matter. The P may have been an integral part of the humic acid structure, or bound through P-iron-humic complexes.
Unreactive P in the <0.45-µm size fraction contributed significant amounts of P on a seasonal basis across all soil types. Unreactive P (<0.45) comprises a range of organic P forms, including inositol phosphates, nucleic acids, nucleotides, phospholipids and sugar phosphates (Turner and Haygarth, 2000), although condensed inorganic P forms, such as pyrophosphates may also be included (Ron Vaz et al., 1993). Maximum concentrations of up to about 70 µg L-1 in the springtime (April–May) (Fig. 4), which mirrored a similar trend for RP (<0.45), were detected over several drainage seasons (Fig. 1). The release of UP (<0.45) during the springtime may be a result of biological processes in the soil. Elevated temperatures and an increase in the availability of carbon substrates in the soil results in a maximum soil microbial biomass size and activity during the springtime (Patra et al., 1990; Lovell et al., 1994). The resulting rapid turnover of microbial biomass through the solution pool may result in the availability of organic P to transfer in drainage water. In addition, wetting and drying cycles during this time may result in a flush of microbial intracellular P through the lysis of desiccated cells by rewetting (Salema et al., 1982; Kieft et al., 1987). The turnover of microbial biomass P has been estimated to be in the region of 25 kg ha-1 yr-1 and is certainly responsible for the release of large amounts of P to the soil (Brookes et al., 1984; Sarathchandra et al., 1989). The biomass released P will be in the form of mobile and labile nucleic acids, phospholipids and sugar phosphates (Bieleski, 1973). As a result, it has a high potential to escape to surface waters if drainage is occurring, where it will provide a bioavailable P source. However, these processes are speculative and require further investigation.
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Conclusions |
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TOPABSTRACTINTRODUCTIONMethodsResultsDiscussionConclusionsREFERENCES |
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ACKNOWLEDGMENTS |
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Received for publication March 22, 1999.
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TOPABSTRACTINTRODUCTIONMethodsResultsDiscussionConclusionsREFERENCES |
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