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TECHNICAL REPORTS
Landscape and Watershed Processes

Uptake and Release of Phosphorus from Overland Flow in a Stream Environment

USDA-ARS, Pasture Systems and Watershed Management Research Unit, Curtin Road, University Park, PA 16802-3702

^* Corresponding author: (richard.mcdowell{at}agresearch.co.nz)

Received for publication April 14, 2002.

	ABSTRACT

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Phosphorus runoff from agricultural fields has been linked to freshwater eutrophication. However, edge-of-field P losses can be modified by benthic sediments during stream flow by physiochemical processes associated with Al, Fe, and Ca, and by biological assimilation. We investigated fluvial P when exposed to stream-bed sediments (top 3 cm) collected from seven sites representing forested and agricultural areas (pasture and cultivated), in a mixed-land-use watershed. Sediment was placed in a 10-m-long, 0.2-m-wide fluvarium to a 3-cm depth and water was recirculated over the sediment at 2 L s^-1 and 5% slope. When overland flow (4 mg dissolved reactive phosphorus [DRP] and 9 mg total phosphorus [TP] L^-1) from manured soils was first recirculated, P uptake was associated with Al and Fe hydrous oxides for sediments from forested areas (pH 5.2–5.4) and by Ca for sediments from agricultural areas (pH 6.5–7.2). A large increase (up to 200%) in readily available P NH₄Cl fraction was noted. After 24 h, DRP concentration in channel flow was related to sediment solution P concentration at which no net sorption or desorption of P occurs (EPC₀) (r² = 0.77), indicating quasi-equilibrium. When fresh water (approximately 0.005 mg P L^-1; mean base flow DRP at seven sites) was recirculated over the sediments for 24 h, P release kinetics followed an exponential function. Microbial biomass P accounted for 34 to 43% of sediment P uptake from manure-rich overland flow. Although abiotic sediment processes played a dominant role in determining P uptake, biotic process are clearly important and both should be considered along with the location and management of landscape inputs for remedial strategies to be effective.

Abbreviations: DRP, dissolved reactive phosphorus • EPC₀, solution P concentration at which no net sorption or desorption of P occurs • P_max, phosphorus sorption maximum derived from Langmuir isotherm • PP, particulate phosphorus • TP, total phosphorus

	INTRODUCTION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
THE LOSS OF PHOSPHORUS (P) in agricultural runoff has been linked to an increased risk of surface water eutrophication (Carpenter et al., 1998; Sharpley, 2000; United States Geological Survey, 1999). However, this risk is only expressed if benthic sediments cannot buffer P inputs to stream P concentrations less than that required for eutrophication. Thus, successful environmental management through the mitigation of P loss requires consideration of not only P inputs, but also the fate of inputs in the freshwater environment. The processes that control P uptake and release in soils and fluvial sediments include a combination of physiochemical processes such as sorption and desorption and coprecipitation with Al and Fe hydroxides and Ca compounds, and biological processes such as assimilation by bacteria, biofilms, and aquatic plants (Mainstone and Parr, 2002). A total ecosystem approach, therefore, requires consideration of P uptake and release in both the landscape and fluvial environments.

To date, much work has been conducted on the movement and chemistry of P within the landscape as a function of inputs such as manures and fertilizers and the potential and actual movement of these inputs and soil-derived P in flowing water. For fluvial systems, the chemistry of P has been extensively studied in experiments using batch incubations where the liquid phase is not replaced (McCallister and Logan, 1978; Ryden et al., 1972; Stone and Mudroch, 1989). However, these are often criticized for their failure to simulate a flowing fluvial system. One alternative is to study P spiraling in situ (e.g., Newbold et al., 1983; Mulholland et al., 1985). Similarly, House and coworkers studied P cycling using a fluvarium where water was continuously cycled over sediments, open to the air, but in a controlled laboratory environment (e.g., House et al., 1995a,b; House and Warwick, 1999; House and Denison, 2000). These authors have been able to describe the kinetics of P uptake by stream sediments using the Elovich and similar diffusion equations and chemical precipitation models. The results showed that P concentration in flow tended toward steady state and was consistent with measurements of the sediment equilibrium P concentration at zero net sorption or desorption (EPC₀).

These initial experiments were conducted with inorganic P (as KH₂PO₄). The behavior of P in fluvial situations is likely to differ in response to natural P inputs (soil P derived from fertilizer and/or manure in dissolved [<0.45 µm] or particulate [>0.45 µm] forms) such as P in overland flow. Thus, our main objective was to establish the behavior of dissolved (inorganic and organic) and particulate forms of P in a simulated fluvial environment when exposed to different benthic sediments.

The relative proportion of inorganic and organic P in dissolved and particulate forms available to loss mechanisms, such as overland flow, differs with the form of P applied. For example, McDowell and Sharpley (2001) showed that much more organic P was present in the soil solution of manured soils than soils that had received the same quantity of P as inorganic fertilizer. Recent work has also shown that soils with recent dairy manure applications conform to the concept of P movement via highly mobile flocs (McDowell and Sharpley, 2002a). Clearly, the form and mobility of land-applied P in inputs is of critical importance once in the fluvial system. Thus, our second objective was to compare the behavior of P species when exposed to different sediments in overland flow from the same soil that had received either dairy or poultry manure.

Recent work is beginning to highlight the importance of biological processes in the uptake and release of P. Several authors have cited sediment bacteria as an important mechanism for the mediation of P dynamics in fluvial systems (e.g., Haggard et al., 1999), while others have found less evidence of their role (e.g., Klotz, 1988). The proportion of P in the microbial biomass can fluctuate in response to, among others, redox conditions, water availability (e.g., wetting drying cycles), and nutrient availability (Gächter and Meyer, 1993; Baldwin et al., 1997, 2000; Golterman, 2001). Conditions within our experimental framework are controlled and facilitate the study of a third objective: investigating the relative size and importance of the microbial P pool in sediments receiving flow from areas of different land use and P inputs (dairy or poultry manure plus soil P in overland flow).

	MATERIALS AND METHODS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Study Site
The study area, WE-38, is a subwatershed of the Manhantango Creek watershed and is characteristic of the mixed agricultural and forested uplands common in the Appalachian Province and Piedmont Plateau of the eastern USA (Fig. 1) . The 7.3-km² watershed has been examined since 1968 in numerous studies, which describe in detail watershed hydrology, topography, geology, and land use (e.g., Pionke et al., 2000). Briefly, WE-38 is located 40 km north of Harrisburg, Pennsylvania, USA, within the Susquehanna River basin, which supplies nearly half of the flow to the Chesapeake Bay. Elevations range from approximately 240 m above mean sea level in the south to approximately 480 m above mean sea level in the north, while slopes range from 3 to 17%. Land use among the 285 delineated fields is dominated by agriculture with 57% of the watershed in cropland, 8% in pasture, and the remaining 35% as deciduous forest. Precipitation is approximately 1100 mm yr^-1 and stream flow approximately 450 mm yr^-1. Approximately 20% of stream flow from the watershed is from overland flow (controlled by variable source area hydrology; Ward, 1984), while the remainder is subsurface flow (Pionke et al., 1999). Residence time of subsurface flow through the watershed is short (1–3 yr) due to high transmissivities and small water storage capacity dictated by fracture, rather than matrix porosity (Pionke et al., 2000).

View larger version (32K): [in this window] [in a new window]   Fig. 1. Location of WE-38 watershed and sampling sites.

The sediments under study were extracted from the northwest corner of the WE-38 watershed and were chosen to represent areas unaffected by agriculture and sites under intensive agricultural management (Fig. 1, Table 1).

Fluvarium and Treatments
All experiments were conducted in a purpose-built fluvarium. Attached to each downslope end of the four 10-m-long by 20-cm-wide by 20-cm-deep troughs (slope angle variable from 0–15%) is a reservoir with a 300-L total capacity (Fig. 2) . Plumbing and a pump for each trough and reservoir allow solution to recirculate over the sediment from the upslope end at rates varying from 1 to 20 L s^-1. It is also possible to alter plumbing so that flow from one trough can be directed into another, thus, up to 40 m of sediment flow path is possible. Attached to each end of the reservoir is a back-flow system, which siphons off a small proportion of flow moving through pipes back into the reservoir to agitate and keep the reservoir solution continually mixed.

View larger version (28K): [in this window] [in a new window]   Fig. 2. Top and side-on views depicting one trough of the fluvarium. Three other troughs sit alongside. Diagram is not to scale.

• Input phase in which manure-rich overland flow is introduced to flow over each sediment.
  
• A flushing phase in which stream flow plus overland flow is replaced by clean water and stream flow reinitiated.
  

For the input phase, in turn, two replicates of each sediment were placed into two troughs of the fluvarium to a depth of approximately 3 cm and the troughs set at a 5% angle (estimated mean slope of the seven sampled sites). Each reservoir was filled with 180 L of tap water (P less than 0.005 mg L^-1 detection limit) and flow pumped over the sediment at 2 L s^-1 for 30 min (mean estimated flow velocity for the seven sites). After this initial equilibration period, 20 L of dairy- or poultry-manure-rich overland flow, both of 4 mg L^-1 dissolved P and 9 mg L^-1 total P, was introduced into the fluvarium reservoir. This overland flow was collected in response to artificial rainfall applied at 6.5 cm h^-1 for 60 min from a Watson silt loam (fine-loamy, mixed, active, mesic Typic Fragiudult) of approximately 75 mg kg^-1 Mehlich 3–extractable P, which had 100 kg P ha^-1 dairy or poultry manure added 3 d prior. Immediately after the introduction of the manure-rich overland flow, a sample was taken at the flume and additional samples taken at the 10-min mark and then every hour for 24 h with an automatic sampler. At the end of the input phase a composite sample was taken from sediment at varying locations along the trough. This sample was divided into two halves; one half was immediately sterilized along with a subsample of untreated sediment with 5M Rad of irradiation from a ⁶⁰Co source for 16 h. All samples were kept wet and in the dark at 278 K until analysis.

For the flushing phase, reservoirs were drained and cleaned with fresh water before filled with 200 L of fresh water (P < 0.005 mg L^-1). Flow was then initiated, again at 2 L s^-1, and samples were taken at the flume at the 10- and 30-min mark, and every hour after the start of flow for four hours.

Soil Analyses
All analyses were conducted in duplicate. A subsample of each wet sediment was air-dried and then oven-dried (378 K) to determine moisture content gravimetrically and for the presentation of all sediment results on an air-dry weight basis. Air-dry sediment was used to determine undispersed particle size distribution among four categories (sand: >63 µm, silt: <63 µm and >2 µm, clay: <2 µm, and fines: the sum of silt and clay fractions) with a hydrometer. Sediment pH was determined with a 1:2.5 sediment to water ratio. Mehlich 3–extractable P was determined by shaking wet sediments (equivalent to 1 g dry wt.) with 10 mL of 0.2 M CH₃COOH, 0.25 M NH₄NO₃, 0.015 M NH₄F, 0.013 M HNO₃, and 0.001 M EDTA for 5 min (Mehlich, 1984), filtering, and analyzing the extract colorimetrically. Sediment exchangeable Ca and cation exchange capacity (CEC) were determined with the ammonium acetate method outlined by Hendershot et al. (1993).

Phosphorus speciation in all samples was determined with a simplified Hieltjes and Lijklema (1980) sequential extraction regime. To 2 g of sediment (dry weight equivalent), 30 mL of 1 M NH₄Cl was added and the suspension shaken overnight (20 h). After shaking, the suspension was centrifuged (2500 x g) for 10 min and the supernatant decanted off, filtered (<0.45 µm), and kept at 278 K until analyzed (within 7 d). This process was repeated twice with 0.1 M NaOH and then once with 1.0 M HCl. These fractions represent readily labile P (NH₄Cl-P), P bound by Al and Fe oxides and humic materials (NaOH-P), and calcium (largely apatite)-bound P (HCl-P), respectively. A subsample of the combined NaOH extract was digested (Taylor, 2000) to give total NaOH-extractable P, and organic P (NaOH-P_org) by difference from inorganic P (NaOH-P). Samples of oven-dried (378 K) soil were also analyzed for total P after aqua regia digestion (4:1 concentrated HCl and HNO₃ mix; Crosland et al., 1995) with inductively coupled plasma mass adsorption spectroscopy. Phosphorus in all neutralized extracts was determined by the method of Murphy and Riley (1962). Microbial biomass P was calculated as the difference between the sum total sequentially extracted P from sterilized and unsterilized samples.

Water Analyses
Water samples were filtered (<0.45 µm) and stored at 278 K in the dark along with unfiltered samples. Within 24 h, each sample was analyzed for dissolved reactive phosphorus (DRP) and within 48 h for total dissolved phosphorus (TDP) after a Kjeldahl digestion (Taylor, 2000). An unfiltered sample was also digested and total phosphorus (TP) measured within 7 d. Dissolved unreactive P was defined as the difference between TDP and DRP and particulate P as TP less TDP, respectively.

Phosphorus Sorption and Desorption
For P sorption parameters, wet sediments (equivalent 1 g dry weight) were mixed with 20 mL of 0.003 M CaCl₂ solutions (equivalent ionic strength of stream water; Klotz, 1988) containing graduated concentrations (0, 1, 2, 4, 10, 20, and 50 mg P mL^-1) of P (as KH₂PO₄) and shaken for 16 h. Samples were then centrifuged and filtered (<0.45 µm), and P was determined colorimetrically. The Langmuir equation was used to obtain estimates of the P sorption maximum (P_max; mg kg^-1). The initial slope of a graph of P sorbed (mg kg^-1) against P remaining in solution (mg L^-1) was used to estimate equilibrium P concentration (EPC₀; mg L^-1) as the solution P concentration at which no net sorption or desorption (0 mg kg^-1) occurred. As time progresses, it is likely that the P_max will change according to changes in the bindings or precipitation of P to sediments (Detenbeck and Brezonik, 1991). As such, the data gained from the Langmuir equation represents a snap-shot or estimate of P_max after 16 h and not the true P_max.

Preliminary data analysis indicated that P fractions during the uptake phase of flow were best fitted to a simple first-order exponential decay equation:

[1]

where C is the concentration of the relative P fraction (mg L^-1), t is the time since the onset of flow, and and ß are constants relating to the initial concentration of P in flow and the exponential decrease in P fraction concentration as a function of time (defined here as a decay parameter), respectively.

Where an exponential equation could not be fitted (e.g., for DRP during the release phase for three sediments) a power function was a good substitute (P < 0.05):

[2]

Statistics
Statistical analyses (t tests, means, and standard errors) were performed with SPSS Version 10.0 (SPSS, 1999). The first-order exponential and power equations were fitted using regression techniques (in SPSS Version 10.0) and the fit assessed through a linear plot of observed versus predicted values giving a r² value. All r² values given are significant at the P < 0.05 level.

	RESULTS AND DISCUSSION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Sediment Characteristics and Response to Flow
General sediment characteristics are given in Table 1 along with the dominant surrounding land use. Sites 2, 3, and 4 were from areas largely in undisturbed forest. At Site 5, a minor component of the surrounding land use (33%) was still in undisturbed forest; however, the majority (66%) was in grazed pasture. In general, site locations increased in their intensity of agricultural land use and decreased in elevation from Sites 1 to 7 (Table 1).

The distribution of P fractions within sediments reflects their location and contributing inputs. For example, sediment from streams draining forested sites (Sites 2, 3, and 4) were more acidic than sediments from sites surrounded by pasture or cultivated land (Sites 1, 5, 6, and 7) (Table 1). As a consequence, relatively more P was held in fractions probably associated with Al- and Fe-hydrous oxides and organic matter (Table 2; 42–44% before treatment), whereas sites surrounded by pasture or cultivated land had sediments with a higher pH and thus, much more P in the HCl fraction (Ca-associated P, largely apatite, Table 2; 48–85% before treatment). On average, among those sites in pasture and cultivated land, the residual fraction (thought not easily plant available) accounted for less sediment total P (30%) than sites in undisturbed forest (36%) (Table 2). This probably reflects the contribution to sediments of overland flow from soils that have been fertilized or manured and contain much more P in plant-available forms than in overland flow from unamended forest soils.

View this table: [in this window] [in a new window]   Table 3. Mean sorption parameters for duplicate sediments after dairy or poultry manure had been introduced into flow.

The distribution of P among different fractions is consistent with the pH-dependant solubility of P species in both soils and sediments. However, while sequential extraction data give information concerning the physiochemical association of P in sediments, caution should be employed if extending the data to also imply bioavailability (Martin et al., 1987). For soils, recent evidence has indicated with solubility equilibrium experiments and confirmed with solid state nuclear magnetic resonance techniques the presence and active nature of Ca-P species in controlling P uptake and release from soil solution (McDowell et al., 2002). For sediments, House and Denison (2000) demonstrated that amorphous iron hydroxide [Fe(PO₄)_x·XOH], vivianite [Fe₃(PO₄)₂·8H₂O], and octacalcium phosphate [Ca₄H(PO₄)₃·2.5H₂O] were active in controlling the EPC₀.

Comparison of P_max values before and after the uptake phase of flow failed to indicate any significant change (paired t test, P > 0.05). However, EPC₀ values did exhibit a significant difference at the P < 0.05 level. This is possibly due to the increase in readily mobile P from manure-rich overland flow (e.g., up to 200% increase in NH₄Cl fractions, Table 2). Comparison of EPC₀ calculated from sorption experiments with the DRP concentration at the end of the uptake phase of flow (24 h) indicates a close relationship between the two parameters, that is, by the end of the experiment the system is approaching quasi-equilibrium (Fig. 3) . The slope of the regression between the two measures (4.3) indicates the difference in sediment to solution ratios used for the batch experiments (0.05 g mL^-1) and found in the fluvarium (0.3 g mL^-1). Although the slope is not equal to the difference in sediment to solution ratios (4.3 vs. 6), the difference is probably due to the nonlinear effect of soil to solution ratio on P release and uptake, which is well documented in the literature (e.g., Hope and Syers, 1976; McDowell and Sharpley, 2001).

View larger version (23K): [in this window] [in a new window]   Fig. 3. Relationship between the dissolved reactive phosphorus (DRP) concentration at the end of the uptake phase of flow and the solution P concentration at which no net sorption or desorption of P occurs (EPC0) determined from batch sorption isotherms. The numbers in parentheses refer to sediment sites (Table 1).

View larger version (30K): [in this window] [in a new window]   Fig. 4. Concentration pattern of P fractions during the uptake phase of flow.

In addition to differences between manure type, differences in likely flocculation may also occur via interaction with the different sediments, as the flow of water over sediments will always generate some quantity of suspended sediment. To test this hypothesis further, a particle size analysis was conducted on undispersed sediments. A significant relationship was generated between the percent sand in the sediments and the decay rate (ß) for PP or TP (Fig. 5) . This indicates that the decay or decrease in PP and TP concentration with time increases with the proportion of sand-sized particles in the sediment, presumably as heavy sand-sized particles drop out of suspension. Such relationships are well known to occur for different particle size fractions during flow (e.g., Govers, 1990).

View larger version (18K): [in this window] [in a new window]   Fig. 5. Relationship between the decay rate parameter (ß) and the percent sand in each sediment. Numbers in parentheses refer to sediment sites (Table 1).

View larger version (25K): [in this window] [in a new window]   Fig. 6. Concentration pattern of P fractions during the release or flushing phase of flow.

The classical view of the role of microbes in P cycling as outlined by Gächter and Meyer (1993) defines microbes as catalysts that facilitate P uptake or release by abiotic processes such as sorption or precipitation–dissolution (e.g., microbial Fe³⁺ oxide reduction; Roden and Edmonds, 1997). As catalysts, the composition and size of the pool is considered constant since they are neither consumed nor produced during the reaction. However, microbes not only house enzymes that serve this purpose, moreover they are distinct organisms, which depend on P as a nutrient for production. As such, the microbial biomass pool could be thought of as a dynamic P pool growing and shrinking with nutrient supply.

The role of microbes in the uptake of P by sediments is a function, but not exclusively, of redox conditions, temperature, and substrate supply (both P and other nutrients such as C). Under oxic conditions the importance of microbes in P uptake has been found to vary greatly. For example, Klotz (1985) and Meyer (1979) found that P removal attributable to biotic processes ranged from 2 to 25%, and concluded that their role was minimal in comparison with abiotic processes. However, Haggard et al. (1999) and Khoshmanesh et al. (1999) found that biotic process accounted for 38 and 45% of sediment P uptake, respectively. These values are similar to sediment P uptake from poultry (34%) and dairy (43%) manure–rich overland flow in our study (Table 6).

Management Perspectives
The location of manure applications within the landscape is central to the potential risk and loss of P into the fluvial system (McDowell et al., 2001). A review of land use and sediment yield by Walling (1999) indicated that fluvial systems with a low sediment delivery ratio exhibit a large buffering capacity and vice versa. Given the large amount of P in fluvial systems that is transported in PP forms, we need to consider the potential for erosion, land use, and sediment dynamics, as well as overall sediment P content, if remedial strategies are to be effective. Assuming a 2 L s^-1 constant flow rate, a particle containing P will be able to move along a channel a maximum of 720 m. Considering the USEPA limit of 0.1 mg TP L^-1 (USEPA, 1994), many of the sediments would not be able to decrease TP concentration below this limit even after 24 h, equivalent to approximately 17.3 km (assuming the same sediment is present along the entire reach). Obviously, consideration of both sediment transport and chemistry play a critical role in both the deposition and uptake of P along a reach. Highly mobile forms of P need to be considered in the long-term environmental management and ecological engineering of streams and waterways. For instance, the forms of P input need to be considered in relation to the distance away from a receiving water body or wetland, where enhanced deposition will later act as a long-term source of sustained P release.

In the short term, our data have highlighted the importance of the microbial biomass in P uptake. In low-order streams (near headwaters) the potential for fluctuating stream heights and thus, wetting and drying cycles, is much greater than for larger streams and rivers. It is well known that wetting and drying cycles can cause P to shift from water-soluble or easily exchangeable pools to more recalcitrant pools (McDowell and Sharpley, 2002b). Perhaps, of greater importance, is the potential for the release of large quantities of P caused by microbial death via desiccation. However, Baldwin et al. (2000) recently showed that repeated wetting and drying cycles select bacteria that are tolerant to periods of desiccation. What remains to be determined is how this is affected by inputs from manure-rich overland flow, how long the selection of tolerant bacteria takes, and thus, how much of a risk wetting–drying cycles pose in the interim.

	CONCLUSIONS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
The input of P from overland flow can form a significant and sustained proportion of P in simulated stream flow. The pattern of P concentration in dissolved and particulate forms follows a similar exponential decay, whereby P was taken up by the sediment or lost from flow by deposition and/or abiotic and biotic processes as time progressed. The rate and magnitude of this decrease were heavily influenced by the chemistry and physical composition of the underlying sediment. It was evident that surrounding land use and subsequent sediment composition determined what fractions accounted for the majority of sediment P uptake. Similarly, the data suggested that the size and importance of the microbial biomass, acting either as a store of P or as a facilitator of abiotic uptake processes, were determined by the management history and likely P inputs of the surrounding land use. As such, it is clear that the location and management of likely P inputs from the landscape need to be considered in conjunction with the behavior of P in the fluvial environment in order for remedial strategies to be effective.

	ACKNOWLEDGMENTS

 
R.W. McDowell gratefully acknowledges the financial support of the USDA-ARS and the New Zealand Foundation of Science and Technology (Contract AGRX002) while conducting the research and preparing this manuscript.

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