Modeling phosphorus in the environment - Chapter 5 potx

105 5 Phosphorus Transport in Streams: Processes and Modeling Considerations Brian E. Haggard University of Arkansas, Fayetteville, AR Andrew N. Sharpley U.S. Department of Agriculture-Agricultural Research Service, University Park, PA CONTENTS 5.1 Introduction 105 5.2 Abiotic and Biotic Processes 106 5.3 Phosphorus Spiraling 109 5.3.1 Determining Phosphorus Spiraling 109 5.3.2 Stream Properties and Phosphorus Spiraling 113 5.4 Algal and Microbial Processes 114 5.5 Stream Sediments and Phosphorus 115 5.5.1 Sediment Source Effects 115 5.5.2 Sediment and Equilibrium Phosphorus Concentrations 117 5.6 Impact of Stream Processes on Eutrophication 120 5.7 Modeling Phosphorus Transport in Stream Channels 122 5.8 Conclusions 124 References 125 5.1 INTRODUCTION Modeling of phosphorus (P) transport from the landscape to aquatic systems represents a number of complex processes, including rainfall–runoff patterns, manure and fertilizer application, soil–water–P interactions, and crop and forage growth. Numerous process- based models are available to accomplish this task, and many of these models are discussed in other chapters of this book. These models integrate large amounts of © 2007 by Taylor & Francis Group, LLC 106 Modeling Phosphorus in the Environment information to simulate catchment-scale P transport from the landscape. They have been reasonably successful in predicting catchment-scale to edge-of-field losses of P and how nutrient and land management affects these losses. However, a major gap in predicting the response of receiving water bodies is the simulation of fluvial processes occurring during P transport from landscape (edge of field) to receiving waters. Although much research relates land management to edge-of-field P losses, little data are available on the fate and transport of this P once in fluvial systems, though this can affect the amount ultimately entering a water body (McDowell et al. 2004). Most importantly, tributaries can act as sinks and sources of P and can influence the effectiveness of watershed best management practices (BMPs) and the response time of impacted water bodies to land remediation (Meals 1992). The modeling of P transport through streams must account for variability in flow, P sources, and in-channel processes (Hanrahan et al. 2001); however, the complexity and spatial variability of these in- stream processes limit the ability of catchment-scale models to simulate P transport in streams (Sharpley et al. 2002). A complex array of physical, abiotic, and biotic processes occurs within the channel and hyporheic zone of streams, and these complex processes, as well as land-use impacts, influence both P concentrations and loads during down- stream transport. Furthermore, these channel processes can greatly influence the short- and long-term impact of P inputs on the degree of eutrophic response of receiving waters. The proportion of P transported during storm and base-flow conditions varies with catchment P sources and the importance of stream-channel and riparian pro- cesses. Though long-term (e.g., annual) inputs and outputs of P to streams may be similar, short-term (e.g., daily and monthly) transport of P from effluent discharges, drainage fields, and other upstream sources may be heavily buffered by stream- channel processes. Streams that have been enriched with P will often act as short- term transient storage zones, releasing dissolved P back into the water column when aqueous concentrations are low (Ekka et al. 2006; Haggard et al. 2005). The release of P from sediment to overlying waters can delay or even mask P loss reductions from catchment-based BMPs (Meals 1992; National Research Council 2000). Despite the importance of these in-stream processes, stream effects on discrete and edge-of-field P inputs are not adequately simulated in many process-based models of catchment P transport. This chapter describes the important fluvial pro- cesses that influence the form and amount of P transported in streams and approaches to model P transfer from the landscape downstream. 5.2 ABIOTIC AND BIOTIC PROCESSES Aquatic systems — in particular stream reaches and networks — may alter the timing, magnitude, and bioavailability of P transport from the landscape further downstream (Meyer et al. 1988; Sonzogni et al. 1982). Several processes — such as sediment sorption and desorption — occur that may influence P transport through aquatic systems (Froelich 1988; Klotz 1988; Taylor and Kunishi 1971), precipitation and dissolution (Fox 1989; House and Donaldson 1986), microbial and algal uptake (Elwood et al. 1981; Hill 1982), and riparian floodplain and wetland retention (Kronvang et al. 1999; Mitsch 1992; Novak et al. 2004). Many of the abiotic processes are influenced or mediated by biota; for example, coprecipitation of © 2007 by Taylor & Francis Group, LLC Phosphorus Transport in Streams: Processes and Modeling Considerations 107 dissolved P with calcite may be biologically mediated during active photosynthesis (Neal 2001). Furthermore, sediment-associated biotia (i.e., microbial organisms) often account for large fractions of P uptake during sediment–P sorption experiments (Haggard et al. 1999). For instance, several studies have found that aquatic biota accounted for 30 to 40% of P uptake and release in wetland and stream sediments (Khoshmanesh et al. 1999; McDowell 2003; McDowell and Sharpley 2003). In contrast, other work has suggested that the microbial community associated with stream sediments played only a small role in P sorption and buffering capacity (Klotz 1988; Meyer 1979). Clearly, the temporary storage of P from these in-channel processes does alter the transport characteristics of P from different landscape positions through streams to a given outlet within a catchment. The importance of these in-channel processes varies with discharge regimes in individual streams, where these abiotic and biotic processes are most likely to be most important during relatively low- or base-flow conditions. Several studies have shown that P retention occurred in stream reaches during low- or base-flow condi- tions (Dorioz et al. 1998; Hill 1982; House and Warwick 1998). Under these flow conditions P may be temporarily retained within a stream channel, but increasing discharge during episodic storm events may resuspend phosphorus and transport particulate phosphorus (PP) further downstream. The resuspension of sediments might also influence dissolved P concentrations in the water column of streams during these higher flow events (House et al. 1995; Koski-Vähälä and Hartikainen 2001). A large fraction of the P load transported downstream may be resuspended from bottom sediments (Svendsen et al. 1995). In contrast, McDowell et al. (2001) observed the opposite trend of dissolved P retention by channel sediments during storm flow and release during base flow. Reflect- ing the dynamic nature and site specificity of in-channel processes, catchment hydrol- ogy, and sources of sediment and P, McDowell et al. (2001) described the mechanisms controlling P release from soil and stream sediments in relation to storm and base flow at four flumes along the channel of a 40 ha, second-order agricultural catchment (Figure 5.1). Base-flow dissolved P concentrations (average of 1997 to 2004) were greater at the catchment outflow (0.042 mg L −1 at flume 1) than at the uppermost flume (0.025 mg L −1 at flume 4), whereas the inverse occurred during storm flow (0.304 mg L −1 at flume 4 and 0.128 mg L −1 at flume 1). Similar trends in total P concentration were observed. However, it is questionable whether short-term pulses in dissolved P have much ecological impact in streams (Humphrey and Stevenson 1992). During storm flow, in-channel decreases in P concentration were indicative of dilution of P originating from a critical source area above the uppermost flume (flume 4), where an area of high soil P intersected an area of high erosion and overland flow potential (Figure 5.1). During base flow, the increase in P concentra- tions downstream was clearly controlled by channel sediments, such that the P sorption maximum of the uppermost flume (flume 4) sediment (532 mg kg −1 ) was far greater than the outlet flume (flume 1) sediment (227 mg kg −1 ). Paralleling these trends, the sediment equilibrium phosphorus concentration (EPC 0 ) of sediment at flume 1 was greater than at flume 4 (0.034 to 0.004 mg L −1 ). Sediment EPC 0 trends were highly correlated to base flow dissolved P concentrations (0.025 mg L −1 at flume 4 and 0.042 mg L −1 at flume 1). © 2007 by Taylor & Francis Group, LLC 108 Modeling Phosphorus in the Environment FIGURE 5.1 The distribution of Mehlich-3 soil P (>100 mg kg −1 ), erosion (>6 mg ha −1 yr −1 ), and mean dissolved P concentration in storm and base flow (mean of 1997 to 2004) and P sorption–desorption properties of channel sediment at four flumes in FD-36, Pennsylvania. (Adapted from R.W. McDowell, A.N. Sharpley, and G. Folmar, J. Environ. Qual. 30, 1587–1595, 2001. With permission.) Watershed boundary Stream and flume location Field boundaries > 100 mg kg –1 Mehlich-3 soil P High erosion, > 6 Mg ha –1 yr –1 0 0.01 0.02 0.03 0.04 0.05 0 0.01 0.02 0.03 0.04 Flume 4 Flume 3 Flume 2 Flume 1 Baseflow dissolved P (mg L –1 ) EPC 0 of stream sediment (mg L –1 ) Baseflow dissolved P EPC 0 0 0.1 0.2 0.3 0.4 0 200 400 600 Stormflow dissolved P (mg L –1 ) P sorption maximum of stream sediment (mg L –1 ) Stormflow dissolved P Stream sediment P sorption max. 4 3 2 1 © 2007 by Taylor & Francis Group, LLC Phosphorus Transport in Streams: Processes and Modeling Considerations 109 In general, P concentrations and loads increase with increasing discharge, especially concentrations and other sediment-bound constituents often display hysteresis in streams, where the chemograph increases rapidly and peaks on the rising limb of the storm-event hydrograph (Richards et al. 2001; Thomas 1988). Thus, P concentrations are often greater on the rising limb when compared with concentrations measured at a similar discharge on the falling limb of the storm-event hydrograph (Sharpley et al. 1976). Although P concentrations are typically greater during this flow regime, the importance of in-stream P retention is minimized because of sediment resuspension and scouring within the channel. Many streams export a very large fraction — greater than 80% — of P loads during episodic storm events (Green and Haggard 2001; Pionke et al. 1996; Richards et al. 2001), whereas Novak et al. (2003) observed that greater than half of dissolved P export occurred during base-flow conditions. A single large storm event may often transport a large portion of the annual P load in many streams. Though stream reaches may show significant P retention during base-flow conditions, P inputs would typically equal outputs on large time scales, such as annual export. However, P deposition on riparian floodplains may be a significant P sink during storm events and stream-bank overflow (Kronvang et al. 1999). 5.3 PHOSPHORUS SPIRALING As P enters streams from discrete and diffuse sources, P cycles from the dissolved inorganic form to the particulate form (abiotic and biotic) and back into the dissolved inorganic form many times while being transported downstream. The number of cycles that may occur within a given stream reach depends on the spiraling length, which is the distance traveled when completing a cycle (Newbold et al. 1981; Stream Solute Workshop 1990). The spiraling length is composed of two basic parts: P uptake length and P turnover length (Stream Solute Workshop 1990). The uptake length, S w , is the average distance dissolved inorganic P, PO 4 , travels downstream before it is removed from the water column through various abiotic and biotic process that occur within a stream channel (Newbold et al. 1981). The turnover length, S p , is the distance traveled in various particulate forms before P is returned to the water column in the dissolved inorganic form (Newbold 1992). The movement of P through a stream reach is tightly coupled with the downstream transport of water, and each P cycle begins downstream from the next, producing a spiral through the stream ecosystem. The use of the spiraling concept and short-term solute injections have been increasingly used to estimate P retention efficiency in streams. 5.3.1 D ETERMINING P HOSPHORUS S PIRALING Stream Solute Workshop (1990), Newbold (1992), and Webster and Ehrman (1996) provide valuable guidance on solute dynamics in streams and the experimental methods to estimate P uptake length. Short-term solute injections of P and a con- servative (hydrologic) tracer are used to estimate P uptake length in a stream reach where PO 4 concentrations downstream from the injection point typically exhibit an exponential decline (Figure 5.2). The solute injection should be at a constant rate © 2007 by Taylor & Francis Group, LLC during storm events (see, e.g., Green and Haggard 2001; Novak et al. 2003). Phosphorus 110 Modeling Phosphorus in the Environment FIGURE 5.2 Dissolved P concentrations in background water samples and water samples collected during a solute injection experiment on September 2, 1999, at Willow Branch, Oklahoma, with a graphical display of P uptake length calculations. (Adapted from B.E. Haggard and D.E. Storm, J. Freshwater Ecol. 18, 557–565, 2003. With permission.) Dissolved P concentration (mg L –1 ) 0.000 0.005 0.010 0.015 0.020 0.025 Solute injection Background Fraction of dissolved P remaining in the water column 0.0 0.2 0.4 0.6 0.8 1.0 Distance from first sampling site downstream of injection point (m) Natural logarithm of the fraction remaining in the water column –1.0 –0.8 –0.6 –0.4 –0.2 0.0 Slope = –0.006 m –1 S w = 167 m 02040 60 80 100 0 20 40 60 80 100 0 20 40 60 80 100 © 2007 by Taylor & Francis Group, LLC Phosphorus Transport in Streams: Processes and Modeling Considerations 111 and should last long enough for the hydrologic tracer, such as chlorine (Cl), to reach a steady concentration at the most downstream end of the stream reach; conductivity might be used to measure when the steady state is reached downstream. Multiple water samples are collected from sites at increasing distances downstream from the injection point, and P concentrations resulting from the injection are corrected for background concentrations at each site and then for losses due to dilution using a hydrologic tracer (see, e.g., Martí and Sabater 1996). The proportion of P remaining in the water column is used in a negative exponential relation to estimate the P uptake rate per unit distance coefficient, k: P x = P o exp(–kx) (5.1) ln(P x /P o ) = –kx (5.2) where P x is the dissolved P concentration (mg L –1 ) at a distance x downstream corrected for dilution and background concentration, P o is the dissolved P concentration (mg L −1 ) corrected for background concentration at the first site downstream from the injection point, and k is the P uptake rate per unit distance coefficient (m −1 ) or slope of the linear relationship in Equation 5.2. The P uptake length, S w (m) is the inverse of the P uptake rate per unit distance coefficient (1/k). Phosphorus uptake length has been measured using radiotracers 32 PO 4 and 33 4 et al. 1981, 1983) and stable PO 4 additions (see, e.g., Butturini and Sabater 1998; 1996; Mulholland et al. 1990; Niyogi et al. 2004; Valett et al. 2002). Phosphorus uptake length generally constitutes greater than 90% of the spiraling length in rather pristine streams (Mulholland et al. 1990; Newbold et al. 1983), and uptake length has been used as an indicator of stream P retention efficiency. The calculation of P turnover length, S p , requires the use of radiotracers and cannot be accomplished with typical short-term solute injections using stable PO 4 . Because the downstream transport of added PO 4 is influenced by stream water velocity, P uptake length, S w , is often strongly correlated to stream discharge or water velocity (see, e.g., Butturini and Sabater 1998; Haggard et al. 2001b; Niyogi et al. 2004). Davis and Minshall (1999) suggested that the mass transfer coefficient, v f , should be used to compare the retention efficiency of different stream reaches and streams; this parameter would help identify underlying abiotic and biotic pro- cesses influencing P retention. The mass transfer coefficient, v f (m s −1 ), is the vertical velocity at which dissolved P is removed from the water column by abiotic and biotic processes occurring within the stream channel (Stream Solute Workshop 1990); the mass transfer coefficient is the P uptake velocity within a stream reach. The mass transfer coefficient is a function of the uptake length, average water depth and average water velocity as v f = hu/S w = Q/(S w w) (5.3) © 2007 by Taylor & Francis Group, LLC PO (see, e.g., Mulholland et al. 1985, 1990; Newbold Davis and Minshall 1999; Haggard et al. 2001b; Macrae et al. 2003; Martí and Sabater 112 Modeling Phosphorus in the Environment where h is the average water depth (m), u is the average water velocity (m s −1 ), Q is the stream discharge (m 3 s −1 ), and w is the average stream width (m). Thus, the mass transfer coefficient reduces the variability observed in P retention associated with stream-water velocity and depth (see, e.g., Davis and Minshall 1999). Doyle et al. (2003) developed the following equation to examine the influence of in-stream processes (i.e., v f ) and hydrogeomorphology (i.e., h and u) on P retention: P x = P o exp(–Lv f /uh) (5.4) where L is the reach length (m). Equation 5.4 is based on mathematical substitution in the preceding equations in this chapter. This approach allowed Doyle et al. (2003) to determine whether a stream reach would be retentive of dissolved P and whether changes in channel form and uptake processes would alter dissolved P retention. Hydrogeomorphology (i.e., h and u) will vary spatially along the longitudinal down- stream gradient and temporally with changes in channel form and discharge of a stream (Doyle et al. 2003). The P uptake rate per unit area, U (mg m −2 s −1 ) may also be calculated as U = (P b Q)/(S w w) = v f P b (5.5) where P b is the average background P concentration (mg L −1 ) measured before the short-term solute injection. Phosphorus uptake rates are the product of the mass transfer coefficient and ambient dissolved P concentrations, and uptake rates will increase with increases in ambient concentrations. Uptake rates may increase linearly with ambient concentrations until a threshold concentration is achieved where abiotic and biotic processes are saturated (Haggard et al. 2005; Niyogi et al. 2004). The aforementioned citations would provide the range of P uptake lengths, velocities, and rates. However, the greatest variability would generally be observed in P uptake length and the least amount in P uptake rates. Short-term solute additions, such as those used to estimate these parameters, measure gross nutrient uptake and assume that the injection duration is short enough to avoid regeneration, that uptake processes are not saturated by the level of enrichment, and that the change in concentration through a reach follows an exponential decay (Webster and Erhman 1996). Several studies have shown that P uptake length measured using stable P additions increases as the level of enrichment increases (i.e., ∆PO 4 ↑) (Haggard and Storm 2003; Hart et al. 1992; Mulholland et al. 1990); the uptake velocity and rate would decrease with increasing levels of enrichment from short-term additions. Despite some constraints in short-term solute injections, this methodology provides an extremely valuable means incorporating both abiotic and biotic processes to evaluate whole-reach measures of P uptake processes within the stream channel. Recently, these parameters have been measured in stream reaches where a natural longitudinal gradient in dissolved P exists (Haggard et al. 2001a, 2005; Martí et al. 2004; Merseburger et al. 2005). Within this context and experimental situation, net uptake length, S w-net , net uptake velocity, v f-net , and net uptake rate, U net , have become the usual terminology. The parameters measure the net retention efficiency because the upstream P sources are continuous (e.g., effluent discharges or other point © 2007 by Taylor & Francis Group, LLC Phosphorus Transport in Streams: Processes and Modeling Considerations 113 sources) and because dissolved P would be retained from and released into the water column continuously by abiotic and biotic processes. Reported net uptake lengths, S w-net , and rates, U net , are an order of magnitude greater than uptake lengths, S w , and rates, U, measured in less disturbed stream ecosystems, whereas net uptake velocities, v f-net , are one to two orders of magnitude less than uptake velocities, v f , (Haggard et al. 2005; Merseburger et al. 2005). These experimental data might be useful in modeling P retention and release through a stream reach where a significant effluent discharge or other point source exists. 5.3.2 STREAM PROPERTIES AND PHOSPHORUS SPIRALING The concept of P spiraling, or the distance traveled downstream by one P molecule as it completes one cycle of uptake and transformations from dissolved to organic forms and back into flow, reveals significant information about the degree to which P changes during transport in rivers (Elwood et al. 1983). Lengths of P spiraling vary from 1 to 1000 m as a function of flow regime, season, bedrock geology, and sediment characteristics (Melack 1995; Munn and Meyer 1990). Similarly, interac- tion of ground water with stream flow within the hyporheic zone can cause increases or decreases in P concentrations depending on stream-bed upwelling or infiltration of P-rich stream flow. The first definitive measurement of P spiraling length was reported for a first- order woodland stream in Tennessee using 32 PO 4 as the tracer. The P spiraling length was 190 m, 165 m of which was in the water, whereas the remainder was in fine particulate organic matter. Other North American workers found that spiraling length ranged from 23 m in November to 99 m in August when the concentration of coarse particulate material was less and P was moving largely in dissolved form (Mulholland et al. 1985). However, during storms the distance traveled by P in particulate material can increase by one or two orders of magnitude above typical uptake lengths (Melack 1995). Differences in geology can have a profound effect on P spiraling length. Munn and Meyer (1990) found that a stream with granite bedrock had a spiraling length of 85 m, whereas in a stream with P-rich volcanic bedrock the spiraling length was 687 m. Factors influencing the P retention and spiraling are notoriously variable along a stream reach. For example, where in-stream geomorphic processes cause size sorting or where sediments are enriched with P due to local contributions of P-rich overland flow, sediments can represent a significant source of P to stream flow, even when inputs from runoff have ceased. Edge-of-field riparian management not only impacts overland flow P removal but also has a strong influence on stream P spiraling (Cooper et al. 1987; Hearne and Howard-Williams 1988). Fencing off pastures from grazing allows palatable aquatic macrophytes to flourish and decreases P spiraling length, whereas riparian afforestation can shade out periphyton and macrophytes, thus increasing spiral length. The placement of riparian afforestation within a catch- ment is therefore expected to have a major influence on downstream surface water quality. Quinn et al. (1993) reported that forested (Pinus radiata) riparian areas at the headwaters of grassland-dominated catchments in New Zealand adversely influenced downstream P water quality: total phosphorus (TP) concentrations from © 2007 by Taylor & Francis Group, LLC 114 Modeling Phosphorus in the Environment grazed and riparian forested catchments of 0.488 and 1.195 mg L −1 , respectively. This was attributed to a lack of ground cover within the riparian zone. Quinn et al. (1993) suggested that tree planting should be sufficiently sparse to allow develop- ment of vegetative ground cover. Fennessy and Cronk (1997) concluded that buffer strips should be located along headwater reaches where most catchment water originates and that storing water high in the catchment decreases downstream ero- sion. In addition, streams and riparian areas in headwaters tend to be narrower than downstream channels, taking less land out of production while maximizing nutrient retention and removal compared to the targeting of larger channels downstream (Fennessy and Cronk 1997). 5.4 ALGAL AND MICROBIAL PROCESSES The rate at which algae and other microbial organisms take up and release dissolved P in the water column represents an important component of P retention within stream reaches, especially during base-flow conditions. A Michaelis–Menton rela- tionship is often used to describe dissolved P uptake as a function of dissolved P concentrations, where P uptake becomes saturated as dissolved P concentrations increase. Several studies have shown that algal and microbial P uptake can become saturated at relatively low dissolved P concentrations ( 0.01 mg PO 4 –P L −1 ) (Bothwell 1985; Mulholland et al. 1990), and this aspect of biotic uptake represents a constraint in typical solute-injection experiments using stable PO 4 additions. The level of P enrichment required to measure differences in dissolved P concentrations between sequential monitoring sites downstream from the injection point often exceeds the concentration where algal and microbial uptake would typically be saturated. Furthermore, typical streams draining agricultural catchments would have dissolved P concentrations that are much greater than 0.01 mg L −1 (see, e.g., Haggard et al. 2001b; Macrae et al. 2003; McDowell et al. 2001; Smith et al. 2005), and effluent dominated streams would have dissolved P concentrations that are one to several orders of magnitude greater than 0.01 mg L −1 (see, e.g., Ekka et al. 2006; Haggard et al. 2005; Martí et al. 2004; Merseberger et al. 2005). It is likely that saturation will influence algal and microbial uptake kinetics in most agricultural and urban catchments where catchment-scale P transport is being simulated by process-based models. Though cellular uptake and growth rates are generally saturated at low concen- trations, maximum biomass accrual in streams often occurs at somewhat greater concentrations (0.015 to 0.050 mg PO 4 –P L −1 ) (Bothwell 1989; Horner et al. 1983; Popova et al. 2006). This range of dissolved P concentrations might be more typical of streams draining agricultural catchments, and therefore algal and microbial uptake likely still plays a significant role in dissolved P retention through stream reaches. However, the importance of algal and microbial P uptake will vary spatially and temporally with dissolved P concentrations in streams and with the uptake kinetics of the algal and microbial community. Dissolved P uptake rates of algae will vary with light, water velocity, temperature, grazing, and time following disturbances within the stream channel (see, e.g., Dodds 2003; Horner and Welch 1981; Mulholland et al. 1994). © 2007 by Taylor & Francis Group, LLC [...]... LLC 116 Modeling Phosphorus in the Environment sink for P (McDowell and Sharpley 2001) As a result of the erosion of subsoils, which are often dominated by silt-sized particles, the predominant form of P transport in these fluvial systems is PP, whereas in sandy catchments most P is transported in dissolved form (Baldwin et al 2002) In a much larger catchment, McDowell et al (2002) examined the processes... Agency, EPA-6 2 5- R-9 2-0 06, Office of Water, Washington, D.C Melack, J.M 19 95 Transport and transformations of P, fluvial and lacustrine ecosystems Pp 2 45 254 in Phosphorus in the Global Environment: Transfers, Cycles, and Management, H Tiessen (Ed.) Chichester, UK: John Wiley & Sons Merseburger, G.C., E Martí, and F Sabater 20 05 Net changes in nutrient concentrations below a point source input in two streams... of flowing water on the quasi-equilibrium established between deposited stream sediment and stream flow Stream-bed sediments were collected from just above the four flumes in the agricultural watershed, FD-36, detailed in Figure 5. 1 The sediments were placed in the fluvarium to a 3-cm depth, and water was recirculated over the sediment at a rate representing the average storm flow during 1997 to 2004 The. .. refinement on the coefficient modelling approach J Environ Qual 30:1738–1746 Hart, B.T., P Freeman, and I.D McKelvie 1992 Whole-stream phosphorus release studies: variation in uptake length with initial phosphorus concentration Hydrobiologia 2 35 236 :57 3 58 4 Hearne, J.W and C Howard-Williams 1988 Modelling nitrate removal by riparian vegetation in a spring-fed stream: the in uence of land-use practices... and velocity Pp 121–134 in Periphyton of Freshwater Ecosystems, R.G Wetzel (Ed.) Dr W Junk, The Hague, The Netherlands Horppila, J and T Kairesalo 1990 A fading recovery: the role of roach (Rutilus rutilus L.) in maintaining high algal productivity and biomass in Lake Vesijarvi, southern Finland Hydrobiologia 200–201: 153 –1 65 House, W.A and F.H Denison 2000 Factors in uencing the measurement of equilibrium... Limited modeling of in- stream processes; lumped into a single stream segment; includes channel erosion and transformation between dissolved and particulate P under low-flow conditions HSPF AnnAGNPS ANSWERS-2000 WEND-P GWLF Notes: These models are discussed in following chapters GWLF = Generalized Watershed Loading Functions when considering fluvial processes The Annualized Agricultural Nonpoint Source... shortcoming of most predictions of nonpoint source impacts on the chemical and biological response of receiving water bodies is the lack of linked watershed and water-body response models There must be a greater linkage of interfacing of these models to translate agricultural management effects on the export of P from watersheds to the point of impact in terms of chemical or biological response of receiving... a source or sink for P; the deposition or resuspension of particulate or sediment-bound P; the in uence of abiotic and biotic controls on P uptake and release in the fluvial system; and the additional complexities of whether stream channels are aggrading or degrading Though modeling landscape processes and export of P from watersheds remains a challenge, as clearly documented in other chapters of this... base-flow value for the years 1997 to 2004 During low or base flow for a further 24 hours, the stream sediment, then charged with P from the previous flow cycle, acted as a source of P to flowing water The enrichment of flowing water dissolved P was related to the EPC0 of the sediment (Figure 5. 6) It is clear that stream sediment and flow characteristics can in uence the concentration of P transport in flowing... significantly in uence the sediment EPC0 determined in the laboratory These equilibration experiments would be best performed on fresh, wet sediments because drying sediments can alter the P sorption characteristics where sediment has a lesser P buffering capacity (Baldwin 1996) Another important consideration is that the chemical composition of the water used in the extraction does in uence the estimation . PA CONTENTS 5. 1 Introduction 1 05 5.2 Abiotic and Biotic Processes 106 5. 3 Phosphorus Spiraling 109 5. 3.1 Determining Phosphorus Spiraling 109 5. 3.2 Stream Properties and Phosphorus Spiraling 113 5. 4. of these models are discussed in other chapters of this book. These models integrate large amounts of © 2007 by Taylor & Francis Group, LLC 106 Modeling Phosphorus in the Environment information. LLC 120 Modeling Phosphorus in the Environment sample (Figure 5. 6). For low- or base-flow conditions, flowing water in the flume was replaced with water at dissolved P concentration of a 0.0 25 mg L −1 ,