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TECHNICAL REPORTS

Waste Management

Field-Scale Evaluation of Phosphorus Leaching in Acid Sandy Soils Receiving Swine Waste

^a USDA-ARS Northwest Irrigation and Soils Research Laboratory, 3793 N. 3600 E., Kimberly, ID 83341-5076
^b Department of Biological and Agricultural Engineering, Box 7625, North Carolina State University, Raleigh, NC 27695-7625
^c Department of Soil Science, Box 7619, North Carolina State Univ., Raleigh, NC 27695-7619

^* Corresponding author (nnelson{at}nwisrl.ars.usda.gov)

Received for publication November 23, 2004.

	ABSTRACT

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Accurate descriptions of P leaching are important because excess P applied to soils can enter surface water via leaching and subsurface transport, thereby negatively impacting water quality. The objectives of this study were to monitor P leaching in soils with a long-term history of waste application, relate soil solution P concentrations to soil P status, and quantify P leaching losses. Soil solution was monitored for 20 mo with samplers installed at 45-, 90-, and 135-cm depths in two pits (1 x 3 x 1.5 m) in Autryville (loamy, siliceous, thermic Arenic Paleudults) and Blanton (loamy, siliceous, semiactive, thermic Grossarenic Paleudults) soils located in a grazed pasture in Sampson County, NC, which had received swine waste for >20 yr. Maximum soil solution P concentrations at 45 cm exceeded 18 mg L^–1 in both soils. Soil solution P concentrations at 90 cm in the Blanton soil were similar to that at 45 cm indicating low P sorption. Soil solution P concentrations at 90 cm in the Autryville soil averaged 0.05 mg L^–1 compared to 10 mg L^–1 at 45 cm. A split-line model related soil solution P concentration to the degree of phosphorus saturation (DPS), identifying a change point at 45% DPS. Phosphorus movement past 45 cm equaled or exceeded surplus P additions for both soils. Long-term waste applications resulted in DPS > 90%, high soil solution P concentrations, and substantial vertical P movement. Phosphorus leaching should be considered when assessing long-term risk of P loss from waste-amended soils.

Abbreviations: Al_ox, oxalate-extractable aluminum • DPS, degree of phosphorus saturation • DRP, dissolved reactive phosphorus • Fe_ox, oxalate-extractable iron • M3-P, Mehlich 3–extractable phosphorus • P_ox, oxalate-extractable phosphorus • P_w, water-extractable phosphorus • TDR, time-domain reflectometry • TP, total phosphorus

	INTRODUCTION

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
UNDER TYPICAL AGRICULTURAL practices soil P concentrations remain less than that required for environmentally significant P leaching (Sims et al., 1998; Lemunyon and Daniel, 2002; Hansen et al., 2002). However, long-term animal waste applications can increase soil P concentrations to the point where P leaching is a potential environmental threat. For example, Ham (1999) documented soil P concentrations in subsoils of waste application fields in the North Carolina Coastal Plain that were 10 times greater than P concentrations in fields not receiving waste. Elevated subsoil P concentrations from repeated waste applications were also documented by Sharpley et al. (1984), King et al. (1990), Mozaffari and Sims (1994), Maguire et al. (2000), and Novak et al. (2002).

Because artificial drainage systems accelerate drainage water transport to surface water bodies and decrease the interaction of drainage water with subsoils, the majority of research concerning P loss through leaching has focused on artificially drained soils. Sims et al. (1998) summarized results of over 20 studies that reported P leaching losses through agricultural drainage systems. Although many of the studies reviewed found low P losses in drainage waters, several studies found P concentrations in excess of 1 mg P L^–1. More recent studies also found elevated P concentrations in drainage water from artificially drained soils (Gardner et al., 2002; Beauchemin et al., 2003).

Although less of an immediate environmental concern, P leaching losses from naturally drained soils can still pose an environmental threat, particularly in sandy soils with low P sorption capacities that receive high P additions from animal waste (Breeuwsma and Silva, 1992; Sims et al., 1998). Early assessment of P leaching losses is advantageous because once a soil profile accumulates high amounts of P, downward movement can continue for many years even when P additions are reduced or eliminated (Breeuwsma and Silva, 1992; Breeuwsma et al., 1995; Schoumans and Groenendijk, 2000).

Recent studies using large intact soil columns or field lysimeters found elevated P concentrations in leachate resulting from inorganic or organic P additions, with mean and maximum P concentrations ranging from 0.01 to 0.8 mg L^–1 and 0.3 to 1.8 mg L^–1, respectively (Leinweber et al., 1999; Turner and Haygarth, 2000; Djodjic et al., 2004; Godlinski et al., 2004). Leinweber et al. (1999) found positive correlations between surface (0–30 cm) soil P status and P concentrations in leachate while other authors did not. Djodjic et al. (2004) concluded that while surface soil P concentrations may be indicative of potential P release, subsoil properties, including P status, have a strong influence on P leaching.

Teflon suction cups or tension lysimeters have been used to monitor P leaching through unsaturated zones in the soil profile at depths ranging from 15 to 180 cm (Andersen, 1994; Tischner et al., 1998; Andersen and Xia, 2001; Brye et al., 2002). These authors found P concentrations ranging from 0.01 to 0.2 mg P L^–1 in the leachate, with the majority of the samples having less than 0.05 mg P L^–1. Some of the monitored soils had elevated soil P concentrations in the surface, but the P status of the soil in the immediate vicinity of the samplers was relatively low (Andersen and Xia, 2001).

Although the risk for P loss through leaching is increased by excess P applications to soils with low P sorption capacities, relatively few field-scale research studies have been conducted in such an environment. Novak et al. (2000) monitored P concentrations in shallow ground water beneath sandy soils receiving swine lagoon effluent that had P accumulations > 100 mg Mehlich3–extractable phosphorus (M3-P) kg^–1 soil down to 110 cm. However, M3-P decreased to near background levels (1–5 mg kg^–1) by a depth of 180 cm and shallow ground water at a depth of 3 m had correspondingly low average P concentrations (<0.05 mg P L^–1).

Research shows that P movement is greatly increased with increasing soil P status (Maguire and Sims, 2002; Sims et al., 2002). However, reported research on P leaching in naturally drained soils has focused on P concentrations in leachate collected at depths where soil P status is well below the P sorption capacity, thus leading to relatively low estimates of P leaching losses (Turner and Haygarth, 2000; Brye et al., 2002; Djodjic et al., 2004; Godlinski et al., 2004). The objectives of this research study were to (i) monitor P leaching throughout the soil profile of soils with a long-term history of waste application and various degrees of P saturation, (ii) relate P leaching losses to soil P status, and (iii) quantify P leaching losses with respect to P additions.

	MATERIALS AND METHODS

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Site Description
The study was conducted on a 21-yr-old farrow-to-wean swine facility in Sampson County, NC, that was managed typical of farms in the Coastal Plain region. Swine waste generated on the farm is treated in an anaerobic lagoon, after which excess lagoon liquid is applied to approximately 19 ha of ‘Coastal’ Bermuda grass [Cynodon dactylon (L.) Pers.] pasture overseeded with triticale (X Triticosecale Wittmack). Pastures are alternately grazed and cut for hay. Lagoon liquid, ranging from 50 to 100 mg P L^–1, is applied at a rate of 300 kg available N ha^–1 yr^–1 throughout the growing season (February through December), which overapplies P by 35 to 50 kg ha^–1 yr^–1 based on average P removal rates of 10 and 50 kg ha^–1 yr^–1 from pasture and hay fields, respectively, receiving swine lagoon effluent (Nelson and Mikkelsen, 2001). It is very likely, however, that pre-regulation waste applications (first 10–15 yr of farm operation) exceeded current application rates.

In situ suction soil solution samplers, constructed from modified polysulfone filter holders equipped with 0.45-µm polyethersulfone membrane filters (Nelson and Mikkelsen, 2005), were installed in Autryville and Blanton soils (Table 1) located in the swine farm's waste application fields (Fig. 1a) . Two pits (1 x 3 x 1.5 m) were excavated in each soil and samplers were installed 40 cm into the sidewalls of the pits so that samplers were beneath undisturbed soil (Fig. 1b). Two samplers were installed at each of three depths (45, 90, and 135 cm) in each pit, for a total of 24 samplers (two soils, two pits, two reps, three depths). Slope in the immediate area of the pits was 0 to 3%, however, the overall slope of the field was 4% with some areas as high as 9%.

View larger version (86K): [in this window] [in a new window]   Fig. 1. (a) Location of pits where soil solution samplers were installed in Autryville and Blanton soils on a 20-yr-old swine farm in Sampson County, NC, and (b) schematic of soil solution samplers as installed at each site in Fig. 1a (time-domain reflectometry [TDR] probes were only installed at Blanton Site I and Autryville Site II).

A continuous vacuum of –12 to –10 kPa was maintained in each sampler by means of a hanging water column as described by Riekerk and Morris (1983), thereby allowing for continuous sample collection when soil water potential was greater than –12 kPa. Soil solution samples were removed from the samplers once every 12 to 20 d from September 2002 through April 2004, for a total of 42 sampling intervals. Samples were stored at <4°C and either frozen or analyzed within 2 wk of collection. Because soil solution samplers were equipped with 0.45-µm polyethersulfone membrane filters, all samples were filtered during collection. Initial sampler testing showed that refiltering samples did not affect P concentration; therefore, samples were not refiltered. Dissolved reactive phosphorus (DRP) concentration in all samples was determined colorimetrically using a Lachat (Milwaukee, WI) QuickChem 8000 autoanalyzer instrument.

Soil Analysis
Before pit excavation, three soil cores (9 cm in diameter by 90 cm deep) were removed from each pit for soil chemical analyses. Soil cores were dissected into 5-, 10-, or 15-cm segments based on soil morphology. Additional soil samples were taken from the side walls of the excavated pits from 90 to 140 cm deep in 10-cm increments. Soil samples were air-dried, ground, and passed through a 2-mm sieve before analyses. Samples were extracted for Mehlich 3–extractable phosphorus (M3-P) (Sims, 2000), water-extractable phosphorus (P_w) (Self-Davis et al., 2000), and oxalate-extractable phosphorus, iron, and aluminum (P_ox, Fe_ox, Al_ox) (Schoumans, 2000). The degree of phosphorus saturation (DPS) was calculated as:

[1]

where P_ox, Al_ox, and Fe_ox are oxalate-extractable elements in mmol kg^–1 soil and is 0.5 for noncalcareous sandy soils (van der Zee and van Riemsdijk, 1988; Schoumans and Groenendijk, 2000). Total phosphorus (TP) was determined by a modified sulfuric acid–hydrogen peroxide digestion procedure (Gasparatos and Haidouti, 2001). Particle size analysis of all samples used for chemical analysis was determined by the pipette method (Gee and Bauder, 1986). Phosphorus concentrations in M3-P, P_w, and TP solutions were determined colorimetrically using the Lachat QuickChem 8000. Phosphorus, Al, and Fe concentration in the oxalate extracts were determined by inductively coupled plasma spectrometry (ICP). A split-line model identifying two slopes (b₀ and b₁) and a change point (d₀) was fit to the average soil solution DRP measured from each soil x pit x depth combination and the corresponding DPS (Eq. [2] and [3]):

[2]

[3]

Intact soil cores 6 cm tall by 5.5 cm in diameter were taken in triplicate from each horizon in each pit for determination of water retention, hydraulic conductivity, and bulk density. Water retention functions were determined with laboratory methods described by Klute (1986) and fit to moisture release curves with a modified form of the van Genuchten equation (Fayer and Simmons, 1995). Saturated hydraulic conductivity was determined using the constant head method as described by Klute and Dirksen (1986). Unsaturated hydraulic conductivity functions were determined using saturated hydraulic conductivity and the van Genuchten method (van Genuchten, 1980). Analysis of variance and contrasts for pair-wise comparisons of soil analysis and soil solution DRP concentrations were performed with SAS proc glm (SAS Institute, 1998). The split-line model and moisture release curves were fit with SAS proc nlin (SAS Institute, 1998).

Determination of Percolate Volume
Percolation volume for each sampling interval was determined using a water balance method, given as:

[4]

where S is the change in storage (cm), P is precipitation (cm), I is irrigation (cm), ET is evapotranspiration, and D is the drainage (cm). A negative value for D indicates water movement is downward through the soil profile and a positive value for D indicates upward flux of the soil water. Runoff was assumed to be negligible because the field was planted in pasture with good ground cover and the surface soil had very high saturated hydraulic conductivities (>35 cm h^–1). Furthermore, there were no visual indications of runoff in the vicinity of the field plots.

Computationally, a water balance was determined separately for each soil layer, given as:

[5]

for the surface soil layer and:

[6]

for subsurface soil layers, where BAL_i is the water balance for layer i, PET_i is the potential evapotranspiration removed from layer i (cm), and other terms are as previously defined. Potential evapotranspiration (PET) was determined by multiplying the reference ET by a crop coefficient. The reference ET was determined with the FAO56 Penman–Monteith combination method on an hourly time step (Allen et al., 1998). Crop growth was divided into three growth stages with three transition periods, where each growth stage had a different crop coefficient (Chamblee et al., 1995; Allen et al., 1998). The fraction of PET removed from each soil layer (PET_i) was determined as:

[7]

where _i are distribution coefficients 0.6, 0.32, 0.05, 0.02, 0.01, and 0 for soil layers 1 through 6, respectively, which were determined with the root distribution function:

[8]

where y is the fraction of roots above depth z and m is the half-depth or depth at which y = 0.5 (Federer et al., 2003). The parameter m was set at 15 cm based on research from Beyrouty et al. (1990), Hays et al. (1991), and Bonachela (1996).

Using Eq. [5] and [6], drainage from each layer was computed as:

[9]

where U_max(i) is the maximum upward flow of soil water that could have occurred during the sampling interval (cm).

If the water table was below the monitoring depth (135 cm), then the maximum upward flux of water into a soil layer [U_max(i)] was determined by estimating average daily flux between the midpoints of consecutive soil layers using Darcy's law:

[10]

where q is soil water flux (cm d^–1), K(h) is the unsaturated hydraulic conductivity as a function of h (cm d^–1), h is the soil water potential at the center of layer i (cm), and z is the distance from the surface to the center of layer i (cm). Moisture release curves were used to determine h from average daily volumetric soil moisture content as measured by TDR. Hydraulic conductivity (K) at each point was determined with the corresponding K(h) relationship. Because K(h) and moisture release relationships were different for each layer, the geometric mean interlayer conductivity (K_m) replaced K(h) in Eq. [10] (Eq. [11], Haverkamp and Vauclin, 1979):

[11]

When the average daily flux calculated by Eq. [10] was negative the maximum upward flux was set to 0.

If the water table was above 135 cm (as determined by volumetric water content equal to saturation), U_max(i) was estimated using a steady state relationship between maximum upward water flux and water table depth (Skaggs, 1981). Daily estimates of U_max(i) were summed for each sampling interval.

Actual evapotranspiration removed from each layer (AET_i) was determined as:

[12]

where D_i is as computed in Eq. [9]. Notice that for D_i U_max(i), AET_i = PET_i.

Phosphorus losses at each sampling depth were calculated by multiplying the P concentration in soil solution by the calculated drainage for each sampling interval. Annual P losses were computed by summing the P losses for each sampling interval through the entire year.

	RESULTS AND DISCUSSION

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Soil Phosphorus Distribution
Both Autryville and Blanton soils had significant P accumulations in the upper 75 cm of the soil profile (Fig. 2) , with M3-P concentrations 40 to 90 times background concentrations for similar soils (Novak et al., 2000). There was a significant soil x depth interaction for P_w, M3-P, P_ox, and TP (p < 0.001). Concentrations of P_w and M3-P followed the same general trend, where the Autryville soil had higher concentrations than the Blanton in the upper 75 cm (p < 0.05), the Blanton soil had higher concentrations from 85 to 115 cm (p < 0.05), and both soils had similar P concentrations at 125 cm and below (p > 0.05). Concentrations of P_ox and TP are higher in the upper 75 cm of the Autryville soil compared to the Blanton (p < 0.05), but were similar below 85 cm (p > 0.05) (Fig. 2b). Although patterns of P accumulation in the two soils are very different, the high soil P concentrations below 50 cm suggest that past excess P applications have saturated the phosphorus sorption capacity of the upper horizons in both soils.

View larger version (34K): [in this window] [in a new window]   Fig. 2. (a) Mehlich 3–extractable phosphorus (M3-P) and water-extractable phosphorus (Pw) concentrations with depth in the Autryville and Blanton soils (95% confidence limits [CL] represent ±43 and ±3 mg kg–1 M3-P and Pw, respectively) and (b) total phosphorus (TP) and oxalate-extractable phosphorus (Pox) concentrations with depth in the Autryville and Blanton soils (95% confidence limits [CL] represent ±74 and ±56 mg kg–1 TP and Pox, respectively).

View larger version (28K): [in this window] [in a new window]   Fig. 3. (a) Oxalate-extractable aluminum (Alox) and iron (Feox) and (b) degree of phosphorus saturation (DPS) with depth in the Autryville and Blanton soils (95% confidence limits [CL] represent ±88 and ±30 mg kg–1 Alox and Feox, respectively, and ±8% DPS).

View larger version (22K): [in this window] [in a new window]   Fig. 4. Monthly precipitation and irrigation during the study compared to the 30-yr average precipitation.

Phosphorus Concentrations in Soil Solution
The average soil solution DRP concentration at 45 cm in the Autryville soil was over two orders of magnitude greater than DRP concentrations at 90 and 135 cm (Table 3). By contrast, soil solution DRP concentrations were elevated in the Blanton soil down to 90 cm, but were below the detection limit (0.03 mg L^–1) at 135 cm. Although the DPS in the Autryville soil is greater than or equal to the DPS in the Blanton soil above 70 cm, soil solution DRP at 90 cm is two orders of magnitude less than the Blanton soil, confirming the importance of considering subsoil properties when estimating the fate and transport of P, as was also noted by Djodjic et al. (2004).

View larger version (38K): [in this window] [in a new window]   Fig. 5. Dissolved reactive phosphorus (DRP) concentrations in soil solution collected at (a) 45 cm in the Autryville soil and at 45 and 90 cm in the Blanton soil and (b) 90 and 135 cm in the Autryville soil and at 135 cm in the Blanton soil.

View larger version (21K): [in this window] [in a new window]   Fig. 6. Dissolved reactive phosphorus (DRP) concentrations in soil solution at 45 cm in the Autryville soil with corresponding drainage rates from the overlying soil layer (0–20 cm; bars) and waste applications.

Split-line models describing two linear relationships on either side of a threshold are used to relate DPS, or other measures of soil P status, to P_w and/or soil solution DRP for use in defining environmental thresholds for P leaching (Hesketh and Brookes, 2000; Maguire and Sims, 2002; Koopmans et al., 2002; Nair et al., 2004). Similarly, a split-line model was fit to the average soil solution DRP measured from each soil x pit x depth combination and the corresponding DPS for the soils in this study. The split-line model identified a change point in the relationship between soil solution DRP and the soil DPS at 45% DPS, after which the DRP concentration increased rapidly with increasing DPS (Fig. 7a) . This change point is slightly lower than the change point of 56% identified by Maguire and Sims (2002), which may be due to our lack of data points between 43 and 95% DPS. The relationship between P_w and DPS was also fit with the split-line model, identifying a change point at 59% DPS (Fig. 7b). The increase in the slope of the P_w vs. DPS relationship after the change point was much less than that in the DRP vs. DPS relationship, as indicated by a smaller b₁ to b₀ ratio (Table 4). The smaller b₁ to b₀ ratio is a result of the smaller soil to solution ratio for the P_w extraction. Koopmans et al. (2002) calculated that soil to solution ratios of 1:0.3 and 1:10 should have b₁ to b₀ ratios of 110 and 7.1, respectively, which are in general agreement with soil to solution and b₁ to b₀ ratios in Table 4. Koopmans et al. (2002) also calculated that the change point should decrease with decreasing soil to solution ratio, which is contrary to what was found in this study. However, this discrepancy could also be due to a lack of data points between 43 and 95% DPS in the DRP vs. DPS relationship, thus suggesting that the change point would have been greater had there been more data points between the extremes. Although change points identified in this study agree with results of Maguire and Sims (2002), they are greater than the change point of 20% DPS found by Nair et al. (2004) and the change point of 25% DPS suggested by research in the Netherlands (Breeuwsma et al., 1995; Schoumans and Groenendijk, 2000).

View larger version (19K): [in this window] [in a new window]   Fig. 7. Relationship between the (a) average soil solution dissolved reactive phosphorus (DRP) concentration and soil degree of phosphorus saturation (DPS) and (b) water-extractable phosphorus (Pw) concentration and DPS for Autryville and Blanton soils fit with corresponding split-line models.

	CONCLUSIONS

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Annual P applications of at least 35 kg excess P ha^–1 yr^–1 for 20 yr on sandy soils resulted in elevated P concentrations in soil and soil solution below the root zone. Substantial P movement occurred in the upper horizons of P saturated soils while soil solution P concentrations remained very low in the lower horizons (DPS < 45%). Therefore, accurate assessments of P leaching in sandy soils must incorporate P status and P sorption capacity of the entire soil profile, not just the surface horizons.

A split-line model adequately described the relationship between average DRP concentration in soil solution and the DPS and identified a change point at 45% DPS. Slope above the change point was 140 times greater than the slope below the change point; therefore, increases in adsorbed P above 45% DPS have 140 times greater impact on soil solution P concentrations compared to soils with less than 45% DPS. A change point of 59% DPS was determined for the P_w vs. DPS relationship, suggesting that more data points for the soil solution DRP vs. DPS relationship may be needed to more precisely identify the change point.

Vertical P movement in excess of surplus P additions was observed in P saturated soil horizons when precipitation was above average, indicating that P desorption contributed up to 50% of P leached past 45 cm in the Autryville soil. Thus, P leaching could continue even in absence of P additions, demonstrating the need to consider long-term leaching losses when calculating potential environmental impacts of P loss from agricultural soils with low P sorption capacity.

Environmental consequences of the observed P leaching losses are determined by eventual P transport to surface water through subsurface lateral flow, which is in turn dependent on future P applications, subsurface flow rates and pathways, and soil characteristics through which lateral flow occurs. Although this analysis is beyond the scope of the present study, the results of the present study show that substantial quantities of P can be leached through soils with low P sorption capacities and further study on the fate and transport of P leached past the root zone is warranted. Future studies should include field evaluations of subsurface P transport and processed-based modeling efforts to predict P sorption, desorption, and transport in the root zone and through subsurface pathways. Modeling efforts would also be useful in identifying the future impacts of current P management strategies compared to alternative best management practices.

	ACKNOWLEDGMENTS

 
The authors thank the Animal and Poultry Waste Management Center at North Carolina State University for providing funding. We also thank Guillermo Ramirez, Peggy Longmire, and Erik Lewis for their assistance in processing and analyzing soil and solution samples.

	NOTES

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Mention of trademarks, proprietary products, or vendors does not constitute a guarantee or warranty of the product by the USDA-ARS and does not imply its approval to the exclusion of other products or vendors that may also be suitable.

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