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Managing natural processes in drainage ditches for nonpoint source phosphorus control.

Drainage ditches are essential to the productivity of a large proportion of agricultural land, enabling farm machinery to operate and crops to grow at critical times. Many states rely on drainage for agricultural production with nearly 37% of arable land in the Midwest requiring drainage (Fausey et al. 1995). For instance, Illinois, Indiana, Iowa, Michigan, Minnesota, Missouri, Ohio, and Wisconsin have drained more than 20.6 x [10.sup.6] ha (51 x [10.sup.6] ac) using both surface and subsurface drainage systems. In Arkansas, Louisiana, and Mississippi total cropland with drainage exceeds 5 x [10.sup.6] ha (12.5 x [10.sup.6] ac), while a further 2.5 x [10.sup.6] ha (6.2 x [10.sup.6] ac) have been identified as needing land drainage (Bengtson et al. 1995). As channels of concentrated flow that connect agricultural fields with downstream surface waters, ditches have the potential to be conduits for contaminants from diffuse and point sources to downstream water bodies. Understanding the processes that operate within drainage ditches is key to their improved management for water quality protection. In the study reported in this paper, sediments were collected from ditches draining agricultural, forested, and mixed agriculture and forest areas to determine which sediment properties influenced the release and uptake of P using an indoor stream channel fluvarium. The results are discussed in terms of how ditch management decisions can be made to decrease phosphorus (P) export from drained areas.

Based upon an understanding of fluvial processes, it is clear that the role of drainage ditches in P transport is more complex than their characterization as simple conduits. Certainly, drainage ditches can act as direct conduits between agricultural fields and surface waters. For instance, ditches short-circuit other flow pathways (e.g., overland, subsurface), converting flow by more tortuous routes into rapid, concentrated flow with a direct connection to natural surface waters. However, drainage ditches can also act as sources or sinks of P.

Once P from the landscape reaches a ditch, in-channel processes partly control its downstream export (figure 1). During high flow, ditches carry eroded sediments and P from agricultural fields to downstream water bodies. Under low flow, ditches may continue to serve as preferential pathways for dissolved P transport. Biological uptake and oxidation/reduction cycles may also affect the P transport characteristics of ditches. In addition, the processes of channel sedimentation, sediment resuspension and sorption/desorption may alternate ditches from sinks to sources of P (figure 1). In the long-term, a net release of P from sediments may occur following the reduction of inputs, significantly delaying ecological recovery from eutrophication (Boers et al. 1998; Haggard et al. 2005).

Processes Controlling In-Ditch Transport of Phosphorus

Two main processes mediate in-ditch changes in P concentrations: (1) sediment geochemical processes and (2) biotic physiological processes. These combine to influence P transport via sediment sorption and desorption (Taylor and Kunishi 1971; Klotz 1988), precipitation and dissolution (House and Donaldson 1986; Fox 1989), microbial and algal uptake (Elwood et al. 1981; Hill 1982), and riparian floodplain and wetland retention (Mitsch 1992; Novak et al. 2004). Many of the abiotic processes are either influenced or mediated by biota, which often account for large fractions of P uptake during sediment P sorption experiments (Haggard et al. 1999; Khoshmanesh et al. 1999). The temporary storage of P by these in-channel processes can alter the amount and form of P as it moves in ditch drains to a watershed outlet.

The importance of in-ditch processes will vary with ditch flow. Several studies have shown that significant P retention occurs in channel reaches during low flow or base flow conditions (Hill 1982; Dorioz et al. 1998; House and Warwick 1998). High discharge during episodic storm events would serve primarily to resuspend finer-sized sediments and transport particulate P further downstream (Svendsen et al. 1995). The resuspension of sediments would also influence dissolved P concentrations in ditches during these high flow events via sorption/desorption dynamics (House et al. 1995; Koski-Vahala et al. 2001).

In contrast, McDowell et al. (2001) observed dissolved P retention during storm flow and release during base flow. McDowell et al. (2001) described 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 watershed. Average base flow dissolved P concentrations from 1997 to 2004 were greater at the watershed outflow (0.042 mg [L.sup.-1] [0.042 ppm]) than in headwaters (0.025 mg [L.sup.-1]), while the inverse occurred during storm flows (0.304 mg [L.sup.-1] at headwaters and 0.128 mg [L.sup.-1] at the outlet). During storm flow, in-channel decreases in P concentration were indicative of dilution of P originating from a critical source area in upper reaches of the watershed, where an area of high soil P intersected an area of high erosion and overland flow potential. During base flow, the increase in P concentrations downstream was controlled by P release from channel sediments.

Sediment sources within fluvial systems include overland flow from the surrounding landscape, bank erosion, and sediments that are resuspended from the channel bottom. In areas with recent gully formation (channel rejuvenation), subsoil material will be a major source of sediment and likely be a net 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 particulate P, whereas in sandy watersheds most P is transported in dissolved form (Baldwin et al. 2002).

High downstream gradients or losses over time in dissolved P have been recorded widely in artificial streams lined with inert substrates, but containing significant growths of periphytic algae (Horner et al. 1991). This indicates that algal uptake and growth can also have an important influence on drainage ditch dissolved P concentrations. Sediment equilibrium P concentrations ([EPC.sub.0]) and uptake by periphyton are variable within ditch channels, reflective of physical hydraulic processes, the management of land adjacent to the ditch, and the form of P occurring in the sediment. This leads to the concept of P spiralling or the distance travelled downstream by one P molecule as it completes one cycle of uptake and transformations from dissolved to particulate forms and release back into the water column (Newbold et al. 1981; Elwood et al. 1983). The use of the spiraling concept in conjunction with short-term solute injections have been increasingly used to estimate P retention efficiency in streams and possibly ditches. Lengths of P spiralling vary from 1 to 1000 m, and are a function of flow regime, season, bedrock geology, and sediment characteristics (Melack 1995; Munn and Meyer 1990). Similarly, interaction of ground water with ditch flow within the hyporheic zone can cause increases or decreases in P concentrations depending upon ditch-bed upwelling or infiltration of P-rich ditch flow.

Ditches as Nutrient Sinks or Sources

The accumulation of agricultural nutrients in ditch channels is driven by sedimentation, sorption, and biological uptake. These processes moderate the impact of edge-of-field nutrient losses on receiving water bodies. Sedimentation occurs across a range of flows and may be promoted by flow control structures, reduction in channel bedslope, and vegetation growing within ditch channels. For instance, eroded soil may be deposited in settling basins or entrained wetlands established in ditches (Braskerud et al. 2000).

Differential sorption properties of sediments help regulate dissolved P concentrations during periods of low flow (Boers et al. 1998; Sallade and Sims 1997a), such that sediments with lower P sorption saturation and greater buffering capacities can exert a disproportionate effect on dissolved P in a water column (Koski-Vahala et al. 2001; Maguire et al. 2002). Even during high flow events, suspended sediments may control dissolved P concentrations through sorption of dissolved P (Sharpley et al. 1981).

Ditches may in some respect function similarly to wetlands. In fact, ditches that are perceived to require clean-out visually resemble emergent wetlands as they might have vegetated sediment deposits and aquatic vegetation in the main channel. Natural and properly constructed artificial wetlands can retain sediments and nutrients and promote denitrification. To maximize nutrient removal, wetlands should have low flow rates and water levels at or below the soil surface (Chambers et al. 1993). For example, a 300 in ditch in China retained 65% of total P over a two-week period (Yin and Lan 1995). Wetlands with outflow channels and disturbed wetlands exhibit decreased P removal capacity (Chambers et al. 1993).

The open flow system and the periodic scouring of sediments from ditches may limit long-term P retention capacity. Rather, ditch sediments may act as long-term net sources of P, particularly following P input reduction. During large runoff events, elevated flows can scour sediments (Bloom 1991). Resuspended sediments likely contain high concentrations of nutrients, as sediment enrichment ratios (i.e., sediment nutrient concentrations relative to nutrient concentrations of soils from which sediments derive) are generally high due to the preferential erosion of fine-textured, P-rich soil particles (Sharpley 1985). During warm weather, stagnant ditch waters can support dissolved P concentrations that are even greater than those seen under the aerobic conditions simulated by laboratory extraction of sediment P with water, as the reduction of ferric iron ([Fe.sup.3+]) to ferrous iron ([Fe.sup.2+]) under anaerobic conditions dissolves P from Fe-phosphate minerals (Sallade and Sims 1997b; Sparks 1999).

In fluvial systems with good hydraulic mixing (such as shallow flowing streams), the availability of P in sediments can be estimated by the equilibrium P concentration ([EPC.sub.0]) at zero net sorption or desorption. Under conditions of low flow (i.e., base flow), a state of quasi-equilibrium exists, such that EP[C.sub.0] will have a major influence on the concentration of dissolved P in solution, whereby P will desorb from sediments if the concentration of P in ditch flow is less than the sediment's [EPC.sub.0], or conversely, P in ditch flow will adsorb to sediments if the concentration is greater than the [EPC.sub.0] of the sediment (Kunishi et al. 1972).

Kleinman et al. (2007) monitored P transport into and from the same drainage ditches on the Delmarva Peninsula as sampled for the present study. They found overland flow from adjacent fields contributed only 3% to 9% of ditch flow annually. Also, overland flow inputs of P to the ditches was less than 25% of annual ditch export of P, and temporal trends in overland and ditch flow P concentrations did not match. Apparently, ditch and groundwater sources of P are the most likely sources and transport pathways controlling P export from these areas (Kleinman et al. 2007). These principles and processes were further evaluated on sediments collected from these Delmarva ditches.

Materials and Methods

Study Location and Background. The study site is located on the University of Maryland Eastern Shore research farm in Princess Anne, Maryland, within the Manokin River Watershed (MRW), which drains into the Chesapeake Bay (figure 2). Prior to its purchase by the University of Maryland Eastern Shore in 1997, the farm had been a commercial broiler production faculty for 20+ years. On average, the farm is only 5.5 m (18 ft) above mean absolute sea level and receives 1,110 mm (44 in) rainfall annually and is representative of the southern Delmarva Peninsula, which has a flat relief and is dominated by poorly drained soils that require extensive open-air drainage ditches to lower the ground water for crop production. Agriculture is dominated by poultry operations, which produce large quantities of manure each year on the Delmarva Peninsula, much of which is land applied (Delmarva Poultry Industry 2005).

The ditches sampled receive water from the poorly drained Othello silt loam (Typic Endoaquult), which is maintained under a corn/wheat/soybean rotation. Most fields are bounded by at least one deep (>2 m [>7 ft]) ditch that is maintained by the local Public Drainage Association and at least two shallow ditches (<2 m [<7 ft]) that are managed by the farm operator. All soils have a high soil test P concentration (average of 420 mg [kg.sup.-1] Mehlich-3 P) due to their long history of receiving poultry litter at rates exceeding annual crop removal of P.

Sediment Collection. Ditch sediments were collected in September 2003 at three locations (figures 2 and 3) from the top 5 cm of the stream bed. Approximately 75 kg (34 lb) were collected at three sites within each ditch, combined and thoroughly mixed. To remove large materials, sediments were passed through a 10-mm screen and stored wet at 4[degrees]C (39[degrees]F) until analysis (within 7 days).

Fluvarium and Treatments. All experiments were conducted in a purpose-built fluvarium (figure 4). Attached to each downslope end of the four 10 m long by 20 cm wide by 20 cm deep (33 ft by 8 in by 8 in) troughs (slope angle variable from 0% to 15%) is a reservoir with a total capacity of 300 L (80 gal) (figure 4). Plumbing and a pump for each trough and reservoir allow solution to re-circulate over the ditch sediment from the upslope end at rates varying from 1 to 20 L [s.sup.-1] (0.04 to 0.71 [ft.sup.3] [s.sup.-1]). It is also possible to alter plumbing so that flow from one trough can be directed into another, providing a flow-path length of up to 40 m (132 ft). Attached to each of the reservoirs 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.

The P uptake and release properties of the ditch sediments were studied during two experimental phases: (1) release phase in which water represents base flow conditions (i.e., low dissolved reactive P [DRP] concentration is introduced to flow over each sediment) and (2) an uptake phase in which clean water was replaced by ditch flow representing the input of surface runoff from heavily manured soils and ditch flow reinitiated.

For the release phase, two replicates of each sediment were placed into two troughs of the fluvarium to a depth of approximately 5 cm (2 in), and the troughs were set at an angle of 5% (estimated mean slope of the three sampled sites). Each reservoir was filled with 180 L (48 gal) of tap water (P less than detection limit of 0.005 mg [L.sup.-1]) and flow pumped over the sediment at a rate of 2 L [s.sup.-1] (0.07 ft3 [s.sup.-1]) for 48 hours (mean estimated flow velocity for the three ditches; Kleinman et al. 2007). Immediately after flow was initiated, a water sample was taken at the flume and additional samples taken after 5, 10, 15, 30, 45, 60, 90, 120 minutes. 3, 4, 5, 6, 8, 10, 12, 14, 16, 20, 24, 28, 32, 36, 40, 44, and 48 hours with an automatic Sigma sampler. All samples were kept wet and in the dark at 4[degrees]C (39[degrees]F) until analysis.

For the uptake phase, 20 L (5.3 gal) of poultry litter-rich surface runoff, of 2.6 mg [L.sup.-1] DRP and 3.5 mg [L.sup.-1] total P (TP), was introduced into the fluvarium reservoir. This surface runoff was collected in response to artificial rainfall applied at 6.5 cm [hr.sup.-1] (2.6 in [hr.sup.-1]) for 60 minutes from an Othello silt loam of c. 420 mg [kg.sup.-1] Mehlich-3 extractable P, which had received 40 to 100 kg P [ha.sup.-1] (44 to 110 lb P [ac.sup.-1]) poultry litter for the last 15 years. Immediately after the introduction of the P-rich surface runoff, a water sample was taken at the flume and additional samples taken as listed above for the release phase. At the end of the uptake phase, a composite sample of about 100 g was taken of sediment at varying locations along the trough. This sample was divided into two halves; one half, along with a subsample of untreated sediment, was immediately sterilized for 16 hours with 5M Rad of gamma irradiation from a [.sup.60.Co] source. All samples were kept wet and in the dark at 4[degrees]C (39[degrees]F) until analysis.

Sediment and Water Analyses. All analyses were conducted in triplicate, with values reported as means. A subsample of each wet sediment was air dried and then oven dried (105[degrees]C [221[degrees]F]) 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 three categories (sand > 63 [micro]m, silt < 63 [micro]m and > 2 [micro]m, clay < 2 [micro]m) using a hydrometer. Soil cation exchange capacity (CEC) was determined with 1M N[H.sub.4]OAc adjusted to pH 7.0 (Hendershot et al. 1993). Organic matter was estimated by loss on ignition and pH using a glass electrode at a 1:2.5 sediment to water ratio (w/w).

Mehlich-3 extractable P was determined by shaking wet sediments (equivalent of 1 g dry weight) with 10 ml of 0.2M C[H.sub.3]COOH, 0.25M N[H.sub.4]N[O.sub.3], 0.015M N[H.sub.4]F, 0.013M HN[O.sub.3], and 0.001M EDTA for 5 minutes (Mehlich 1984); filtering (<0.45 [micro]m); and analyzing the extract colorimetrically (Murphy and Riley 1962). The inorganic P (IP) concentration of ditch sediments was estimated by shaking wet sediments (equivalent of 0.5 g dry weight) with 0.5 M [H.sub.2]S[O.sub.4] end-over-end for 16 hours (Walker and Adams 1958). Ammonium oxalate extractable Al and Fe was determined by shaking 0.25 g soil in 10 mL of acid oxalate solution (0.1M (N[H.sub.4])[.sub.2][C.sub.2][O.sub.4] [H.sub.2]O and 0.1M [H.sub.2][C.sub.2][O.sub.4] 2[H.sub.2]O) for 4 hours in the dark (McKeague and Day 1966). Oxalate extractable Al and Fe was measured by ICP on filtered extracts.

Fluvarium water samples were filtered (<0.45 [micro]m) and stored at 4[degrees]C (39[degrees]F) in the dark along with unfiltered samples. Within 24 hours, each sample was analyzed for DRP and within 48 hours, for total dissolved reactive P (TDP) following acid persulfate digestion and filtration (Whatman No. 40 filter paper) (Patton and Kryskalla 2003). Phosphorus in all filtrates and neutralized digests was measured by the colorimetric method of Murphy and Riley (1962). An unfiltered sample was also digested and TP measured within 7 days.

Phosphorus Sorption and Desorption. For P sorption parameters, wet ditch sediments (equivalent 1 g dry weight) were mixed with 20 mL of 0.003M Ca[Cl.sub.2] solutions (equivalent ionic strength of stream water [Klotz 1988]) containing graduated concentrations (0, 1, 2, 4, 10, 20 and 50 mg P [mL.sup.-1]) of P (as K[H.sub.2]P[O.sub.4]) and shaken for 24 hours. Samples were then centrifuged, filtered (<0.45 [micro]m), and solution P concentration (C; mg [L.sup.-1]) was determined colorimetrically (Murphy and Riley 1962). The amount of P sorbed (X; mg [kg.sup.-1]) is the difference between P added and P remaining in solution. Using the Langmuir sorption equation, soil P sorption maximum ([P.sub.max]; mg [kg.sup.-1]) was calculated as the reciprocal of the slope of the plot C/X vs. C and binding energy as slope/intercept of the same plot (Syers et al. 1973). The initial linear slope of a graph of P sorbed against P remaining in solution (mg [L.sup.-1]) was used to estimate equilibrium P concentration (EP[C.sub.0]; mg [L.sup.-1]) as the solution P concentration at which no net sorption or desorption (0 mg [kg.sup.-1]) occurred. Langmuir P sorption isotherms were also constructed on irradiated wet ditch sediments as described above. In all cases, sorption of P by sediments was described by Langmuir-based isotherms with an [R.sup.2] of 0.84 to 0.98.

Statistics. Statistical analyses (t-tests, means, and standard errors) were performed with SPSS v. 10.0 (SPSS 1999). All [R.sup.2] values given are significant at the p < 0.05 level.

Results and Discussion

Ditch Sediment Properties. The selective erosion of fine particulates from cultivated agricultural lands increased the silt and clay content of ditch sediments draining agricultural land uses compared with mixed and forest land uses (table 1). Associated with this was an appreciably greater Mehlich-3 P concentration of ditch sediments from agricultural land (107 mg [kg.sup.-1]) than other land uses (29 and 44 mg [kg.sup.-1]). Other studies have shown that sediment P concentrations were greater in fluvial channels with adjacent agricultural land use (McDowell et al. 2002) or predominantly agricultural watersheds (Popova et al. 2006). Sediment organic matter contents were greater in ditches draining agricultural land (77 g [kg.sup.-1]), possibly a function of greater aquatic plant and macrophyte production, as compared with ditches draining mixed (7 g [kg.sup.-1]) and forested (2 g [kg.sup.-1]) lands (table 1).

As expected, the mean [EPC.sub.0] of ditch sediments from agricultural land was more than 10-fold greater than for the other ditches (table 2), reflecting the direct input of P from agricultural fields to these ditches. As EP[C.sub.0] increases, the potential for P release from sediments increases, such that sediments in agricultural ditches will be a source of P as long as the dissolved P concentration of ditch water is <0.38 mg [L.sup.-1] (table 2). Conversely, sediment in the mixed and forest ditches will more likely be a sink for P leaving the agricultural fields. However, the agricultural ditch sediments have a greater capacity to sorb P (362 mg [kg.sup.-1]) than the other ditches (194 and 277 mg [kg.sup.-1]) because of a greater proportion of clay-sized material and more oxalate soluble Al and Fe, which is known to correlate well with P sorption (tables 1 and 2). Further, sorbed P is held more tightly in agricultural ditch than mixed and forest ditch sediments.

Release and Uptake of Phosphorus from Ditch Sediments. Under base flow conditions (initial DRP of 0.005 mg [L.sup.-1] reflecting field conditions), sediments acted as a source of P to overlying water (figure 5A). Due to an appreciably greater Mehlich-3 and [EPC.sub.0] concentration of sediment from agricultural than mixed and forest ditches, a greater concentration of P in ditch flow was attained after 48 hours of flow (figure 5A). For agricultural ditch sediments, a quasi-equilibrium [DRC.sub.0] concentration in fluvarium water of 0.374 mg [L.sup.-1] was maintained at 48 hours. For mixed and forested ditch sediments this concentration was 0.012 and 0.006 mg [L.sup.-1], respectively (figure 5A).

The quasi-equilibrium DRP concentration of fluvarium water was related to the mean [EPC.sub.0] of ditch sediment (figure 6). When data from a similar experiment using stream sediments from an agricultural stream draining an agricultural watershed in south central Pennsylvania were included in this assessment (FD-36 [McDowell and Sharpley 2003]), the relationship between DRP and EP[C.sub.0] was highly significant (p < 0.01; [R.sup.2] of 0.97; figure 6). This suggests that the natural processes influencing P release from deposited sediments is similar between ditches and stream channels. Recent studies have shown that the relationship between stream water dissolved P and benthic sediment Mehlich-3 P concentration was similar to that in soil-water systems (Haggard et al. forthcoming), suggesting this natural process is similar for soils, flows in channels, and stream sediments.

The initial DRP concentration of fluvarium water was then set at 2.6 mg [L.sup.-1], reflecting input from adjacent surface runoff plots (Kleinman et al. 2007). Agricultural ditch sediments were able to sorb or sequester a greater amount of P from ditch flow than sediments from mixed and forest ditches (figure 5B). After 48 hours of flow, agricultural ditch sediments had decreased DRP concentrations from 2.6 to 0.194 mg [L.sup.-1]. Under the same conditions, mixed and forest ditch sediments decreased DRP to only 0.858 and 2.096 mg [L.sup.-1], respectively (figure 5B).

The relative ability of sediments to sequester DRP in ditch flow was closely related to the P sorption maximum of bottom sediments (figure 7). As for the P release phase, when stream channel sediments from FD-36 were included, there was a highly significant relationship between DRP and sediment P sorption maximum ([R.sup.2] of 0.92). As the P sorption maximum of channel sediments increased, the capacity of the sediments to sequester DRP from the overlying water increased where sediments acted as a sink for P. As for the release phase (figure 6), the ability to sequester P was similar for sediments from ditches and natural stream channels (figure 7). This is important in terms of extending our knowledge of sediment-water-P processes occurring in stream channels to help understand certain natural biogeochemical P processes taking place in constructed drainage ditches.

Biological Processes and Sediment P Uptake and Release. Phosphorus sorption properties were determined before and after gamma irradiation to determine the influence of biotic or microbial processes on sediment P uptake and release. Irradiation increased the [EPC.sub.0] of agricultural ditch sediments from 0.383 to 0.535 mg [kg.sup.-1] (40% increase), while [EPC.sub.0] increased 13% and 12% for mixed and forested land use ditches, respectively (table 2). In contrast to P release potential, P sorption maximum decreased from 362 to 304 mg [kg.sup.-1] (18% decrease) and P binding energy from 1.217 to 0.551 L [kg.sup.-1] (55% decrease) with irradiation of agricultural ditch sediments. Sorption properties of sediment from the mixed land use and forest ditches were less affected by irradiation (table 2).

The amount of inorganic P sequestered by sediments during the uptake phase was greater for ditches draining agricultural land (57 mg [kg.sup.-1]) than those draining mixed (23 mg [kg.sup.-1]) and forest lands (18 mg [kg.sup.-1]; table 3). Microbial P determined as the difference in IP between irradiated and non-irradiated sediments was appreciably greater in the agricultural ditch sediment (24 mg [kg.sup.-1]) than the other sediments (4 and 7 mg [kg.sup.-1]) (table 3). For sediments in the ditch draining the agricultural field, microbial processes accounted for 42% of P uptake, decreasing to 22% in the forested ditch sediments.

Clearly, biotic processes play an important role in the uptake and release of P by ditch sediments. In the agricultural ditch, it appears that biotic processes accounted for about 40% of the observed P release ([EPC.sub.0]) and uptake (IP increase). For ditches draining the other land uses, biotic processes were less important and were about 10% for P release and 25% for P uptake. These proportions of biotic and aboitic controls on P release and uptake are consistent with similar studies on P transport in stream channels. For instance, Khoshmanesh et al. (1999) and McDowell and Sharpley (2003) found that aquatic biota accounted for 30% to 40% of sediment P release and uptake in a wetland and stream sediment, respectively (Haggard et al. 1999). 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 (Meyer 1979; Klotz 1988). Evidently, the temporary storage of P by these in-channel processes can alter P transport characteristics in drainage ditches as well as streams.

Best Management Practices

Several important conservation or best management practices (BMPs) can reduce nutrient losses and transport through drainage ditches.

Controlled Drainage. Water control structures at the final point of subsurface drainage outlets can be used to regulate water depth in the ditch, field-water table depth, and water outflow. Water level can be lowered to allow access for farm machinery at critical times. The water level can be raised when desirable, resulting in several beneficial effects, such as (1) providing water storage in the field for use by crops during dry periods; (2) reducing the amount of drainage water, which decreases nutrient export load; (3) increasing denitrification, which reduces N[O.sub.3]-N loss; and (4) increasing sediment and particulate P retention. Negative effects of water control structures include possibly increasing DRP loss from sediments under anaerobic conditions and maintenance costs for outlet pipes.

Sediment Removal. Surface runoff preferentially erodes small soil particles that are rich in P. These P-rich particles often naturally deposit in drainage ditches, where they can continue to release P, as well as physically reduce water flow through the ditches. Therefore, ditches often require periodic clean-outs to maintain flow capacity for adequate drainage. Generally, ditches are dredged using a backhoe, and sediments are deposited on ditch banks and adjacent field edges. Because of its high organic matter and nutrient content, ditch sediments may act as a significant nutrient source, resulting in a positive effect on soil structure and water-holding capacity. In some cases, it may be preferable to move these ditch sediments to areas deficient in soil test P, or at least further away from the ditch or stream channel to areas with low risk for surface runoff and erosion.

Clean-outs may represent a severe ecological disturbance and may function to either increase or decrease nutrient loads. Specifically, significant flow events following disturbance may generate disproportionately large sediment loads. The newly exposed soil present in a ditch may not have the same sediment and nutrient buffer capacity as the original soil. Clean-outs may also function to remove enriched sediments from the flow system and may expose non-enriched subsurface sediments that could act as a new nutrient sink. However, there are few studies evaluating the effect of clean-outs as a BMP to minimize the potential for P losses, enabling profitable crop production on soils where with near-surface water tables. Therefore, more studies are needed to assess whether clean-out can be used as a BMP.

Chemical Treatment. If drainage ditches act as a P source to drainage waters, then another alternative may be to consider chemical treatment of the bottom sediments to decrease sediment [EPC.sub.0] and increase P sorption capacity (Haggard et al. 2004; Smith et al. 2005). Chemical amendments have been successful at decreasing P release under aerobic and anaerobic conditions from sediments in lakes and reservoirs (Cooke et al. 1993). The potential of remedial management practice is described in more detail by Kleinman et al. (2007).

Vegetated Buffers. Buffers can retain nutrients and sediment, decreasing inputs into drainage ditches. Also, grass-vegetated ditch bottoms and sides will reduce in-ditch erosion and may increase biotic uptake of P. Harvesting vegetated buffers can remove significant amounts of P from the ditch. However, when nutrient inputs to drainage ditches are decreased, we must consider the potential for these ditches to become nutrient sources to drainage waters.

Two-Stage Drainage Ditches. The use of two-stage or multi-level ditches is being advocated in some areas to make these systems more self-sustaining (Powell et al. 2007). Flows that exceed the capacity of the inset channel spread out across grassed benches that serve as small active flood-plains. As these out-of-channel flows are shallower and have lower velocities, when compared to a single-stage system, they have a reduced erosive potential (Jayakaran et al. 2005). Stability of the ditch bank may also be improved where the toe of the ditch bank meets the bench rather than the ditch bottom. Overall bank height is effectively reduced and shear stresses on the toe of the bank are less. Dimensions of the low-flow channel can be empirically determined based on regional studies similar to those that are conducted for natural streams. A multi-level ditch has the potential to create and maintain better vegetative and ecological habitat. Also, multi-level ditches may even improve overall water quality and enhance P retention and assimilation (Ward et al. 2004).

Summary and Conclusions

Several interdependent in-ditch processes influence the amounts and forms of P transported from edge-of-field agricultural sources to the point of impact (i.e., river, lake, reservoir, and in this case the Chesapeake Bay). These naturally occurring processes will be critical in defining agricultural source management and in determining eutrophic response of downstream waters. Without information on the direction and magnitude of change in P transport in drainage ditch systems, BMPs will not efficiently remediate against impairment of receiving waters.


Mention of trade names does not imply endorsement by the USDA or the Soil and Water Conservation Society.


Baldwin. D.S., A.M. Mitchell, and J.M. Olley. 2002. Pollutant-sediment interactions: Sorption, reactivity and transport of phosphorus. In Agriculture, Hydrology and Water Quality, ed. Haygarth, P.M. and S.C., Jarvis. 265-280. Oxford, UK: CABI International.

Bengtson, R.L., C.E. Carter, J.L. Fouss, L.M. Southwick, and G.H. Willis. 1995. Agricultural drainage and water-quality in the Mississippi Delta. Journal of Irrigation and Drainage Engineering-ASCE 121:292-295.

Bloom, A.L. 1991. Geomorphology: A Systematic Analysis of Late Cenozoic Landforms. Englewood Cliffs, NJ: Prentice Hall, Inc.

Boers, P.C.M., W. Van Raaphorst, and D.T. Van der Molen. 1998. Phosphorus retention in sediments. Water Science and Technology 37:31-39.

Braskerud, B.C., H. Lundekvam, and T. Krogstad. 2000. The impact of hydraulic load and aggregation on sedimentation of soil particles in constructed wetlands. Journal of Environmental Quality 29:2013-2020.

Chambers J.M., T.J. Wrigley, and A.J. McComb. 1993. The potential use of wetlands to reduce phosphorus export from agricultural catchments. Fertilizer Research 36:157-164.

Cooke, D.G., E.B. Welch, S.A. Peterson, and P.R. Newroth. 1993. Restoration and management of lakes and reservoirs. 2nd ed. Boca Raton. FL: CRC Press.

Delmarva Poultry Industry. 2005. Delmarva and U.S. Facts.

Dorioz, J.M., E.A. Cassell, A. Orand, and K.G. Eisenman. 1998, Phosphorus storage, transport and export dynamics in the Foron River watershed. Hydrological Processes 12:285-309.

Elwood. J.W., J.D. Newbold, A.F. Trimble, and R.W. Stark. 1981. The limiting role of phosphorus in a wood-land stream ecosystem: effects of P enrichment on leak decomposition and primary producers. Ecology 62:146-158.

Elwood, J.W., J.D. Newbold, R.V. O'Neil, and W.Van Winkle. 1983. Resource spiraling: An operational paradigm for analyzing lotic ecosystems. In Dynamics of Lotic Ecosystems, ed. Fontaine III.T.D, and S.M. Bartell. 3-27. Ann Arbor. MI: Ann Arbor Science.

Evans, R., J.W. Gilliam, and W. Skaggs. 1996. Controlled drainage management guidelines for improving drainage water quality. NC Cooperative Extension Service, Publication AG443.

Fausey, N.R., L.C. Brown, H.W. Belcher, and R.S. Kanwar. 1995. Drainage and water quality in great-lakes and corn-belt states. Journal of Irrigation and Drainage Engineering-ASCE 121:283-288.

Fox, L.E. 1989. A model of inorganic control of phosphate concentrations in river waters. Geochimica et Cosmochimica Acta 53:417-428.

Haggard, B.E., D.R. Smith, and K.R. Brye. Forthcoming. Effect of sediment equilibrium P concentration and Mehlich-3 P on dissolved P concentrations at selected streams. Journal of Environmental Quality.

Haggard. B.E., E.H. Stanley, and R. Hyler. 1999. Sediment-phosphorus relationships in three northcentral Oklahoma streams. Transactions of the ASAE 42:1709-1714.

Haggard. B.E., E.H. Stanely. and D.E. Storm. 2005. Nutrient retention in a point source enriched stream. Journal of the North American Benthological Society 24:29-47.

Haggard, B.E., B.A. Ekka, M.D. Matlock, and I. Chaubey. 2004. Phosphate equilibrium between stream sediments and water: Potential effects of chemical amendments. Transactions of the ASAE 47:1113-1118.

Hendershot, W.H., H. Lalande, and M. Duquete. 1993. Ion exchange and exchangeable cations. In Soil Sampling and Methods of Analysis, ed. Carter, M.R., 167-176. Boca Raton, FL: Lewis Publishers.

Hill, A.R. 1982. Phosphorus and major cation mass balances for two rivers during summer low flows. Freshwater Biology 12:293-304.

Horner, R.R., and E.B. Welch. 1981. Stream periphyton development in relation to current velocity and nutrients. Canadian Journal of Fisheries and Aquatic Sciences 38:449-457.

House, W.A., and L. Donaldson, 1986. Adsorption and coprecipitation of phosphate with calcite. Journal of Colloid and Interface Science 112:309-324.

House, W.A., and M.S. Warwick. 1998, A mass-balance approach to quantifying the importance of in-stream processes during nutrient transport in a large river catchment. Science of the Total Environment 210/211:139-152.

House, W.A., F.H. Denison. and P.D. Armitage. 1995. Comparison of the uptake of inorganic phosphorus to a suspended and stream bed-sediment. Water Research 29:767-779.

Jayakaran, A., D. Mecklenburg, A. Ward, L. Brown, and A. Weekes. 2005. Formation of fluvial benches in headwater channels in the Midwestern region of the USA. International Journal of Agricultural Engineering 14:193-208.

Khoshmanesh, A., B.T. Hart, A. Duncan, and R. Beckett. 1999. Biotic uptake and release of phosphorus by a wetland sediment. Environmental Technology 29:85-91.

Kleinman, P.J.A., A.L. Allen, B.A. Necdelman, A.N. Sharpley, P.A. Vadas. L.S. Saporito, G.J. Folmar, and R.B. Bryant. 2007. Dynamics of phosphorus transfers from heavily manured Coastal Plain soils to drainage ditches. Journal of Soil and Water Conservation 62(4):225-235.

Klotz, R.L. 1988. Sediment control of soluble reactive phosphorus in Hoxie Gorge Creek. New York. Canadian Journal of Fisheries and Aquatic Sciences 45:2026-2034.

Koski-Vahala, J., H. Hartikainan, and P. Tallberg. 2001. Phosphorus mobilization from various sediment pools in response to increased pH and silicate concentration. Journal of Environmental Quality 30:546-552.

Kunishi, H.M., A.W. Taylor, W.R. Heald, W.J. Gburek, and R.N. Weaver. 1972. Phosphate movement from an agricultural watershed during two rainfall periods. Journal of Agricultural and Food Chemistry 20:900-905.

McDowell, R.W., and A.N. Sharpley. 2003. The uptake and release of phosphorus from overland flow in a stream environment. Journal of Environmental Quality 32:937-948.

McDowell, R.W., and A.N. Sharpley. 2001. Approximating phosphorus release from soils to surface runoff and subsurface drainage. Journal of Environmental Quality 30:508-520.

McDowell, R.W., A.N. Sharpley, and A.T. Chalmers. 2002. Chemical characterisation of fluvial sediment: The Winooski River, Vermont. Ecological Engineering 18:477-487.

McDowell, R.W., A. Sharpley, and G. Folmar. 2001. Phosphorus export from an agricultural watershed: linking source and transport mechanisms. Journal of Environmental Quality 30: 1587-1595.

McKeague. J.A., and J.H. Day. 1966. Dithionite and oxalate-cxtractable Fe and Al as aids in differentiating various classes of soils. Journal of Soil Science 46:13-22.

Maguire, R.O.A.C. Edwards. J.T. Sims, P.J.A. Kleinman, and A.N. Sharpley. 2002. Effect of mixing soil aggregates on the phosphorus concentration of surface waters. Journal of Environmental Quality 31:1294-1299.

Mehlich, A. 1984. Mehlich 3 soil test extractant: A modification of Mehlich 2 extractant. Communications in Soil Science and Plant Analysis 15:1409-1416.

Melack, J.M. 1995. Transport and transformations of P. fluvial and lacustrine ecosystems. In Phosphorus in the Global Environment: Transfers. Cycles and Management, ed. H. Tiessen, 245-254. Chichester, UK: SCOPE 54 and John Wiley and Sons.

Meyer, J.L. 1979. The role of sediments and bryophytes in phosphorus dynamics in a headwater stream ecosystem. Limnology and Oceanography 24:365-375.

Mitsch, W.J. 1992. Landscape design and the role of created, restored and natural riparian wetlands in controlling nonpoint source pollution. Ecological Engineering 1:27-47.

Munn, N.L., and J.L. Meyer. 1990. Habitat-specific solute retention in two small streams: An intersite comparison. Ecology 71:2069-2082.

Murphy, J., and J.P. Riley. 1962. A modified single solution method for determination of phosphate in natural waters. Analytica Chimica Acta 27:31-36.

Newbold, J.D., J.W. Elwood, R.V. O'Neill, and W. van Winkle. 1981. Measuring nutrient spiraling in streams. Canadian Journal of Fisheries and Aquatic Sciences 38:860-863.

Novak, J.M., K.C. Stone, A.A. Szogi, D.W. Watts, and M.H. Johnson. 2004. Dissolved phosphorus retention and release from a coastal plain in-stream wetland. Journal of Environmental Quality 33:394-401.

Patton, C.J., and J.R. Kryskalla. 2003. Methods of analysis by the United States Geological Survey National Water Quality Laboratory: Evaluation of acid persulfate digestion as an alternative to Kjeldahl digestion for the determination of total and dissolved nitrogen and phosphorus in water. United States Geological Survey. Water Resources Investigations Report 03-4174. Denver, CO: U.S. Geological Survey, Branch of Information Services. Federal Center.

Popova, Y.A., V.G. Keyworth, B.E. Haggard, D.E. Storm, R.A. Lynch, and M.E. Payton. 2006.Stream nutrient limitation and sediment interactions in the Eucha-Spavinaw Basin. Journal of Soil and Water Conservation 61: 105-115.

Powell, G.E., A.D. Ward, D.E. Mecklenburg, and A.D. Jayakaran, 2007. Two-stage channel systems Part 1, a practical approach for sizing agricultural ditches Journal of Soil and Water Conservation. 62(4):277-286.

Sallade, Y.E., and J.T. Sims. 1997a. Phosphorus transformations in the sediments of Delaware's agricultural drainageways. I. Phosphorus forms and sorption. Journal of Environmental Quality 26:1571-1579.

Sallade, Y.E., and J.T. Sims. 1997b. Phosphorus transformations in the sediments of Delaware's agricultural drainageways. II. Effects of reducing conditions on phosphorus release. Journal of Environmental Quality 26:1579-1588.

Sharpley, A.N. 1985. The selective erosion of plant nutrients in runoff. Soil Science Society of America Journal 49:1527-1534.

Sharpley, A.N., T. Krogstad, R.W. McDowell, and P.J.A. Kleimman. 2003. Phosphorus transport in riverine systems. Encyclopedia of Water Science. 10.1081/E-EWS-120020503.

Sharpley, A.N., R.G. Menzel, S.J. Smith, E.D. Rhoades, and A.E. Olness. 1981. The sorption of soluble phosphorus by soil material during transport in runoff from cropped and grassed watersheds. Journal of Environmental Quality 10:211-215.

Smith, D.R., B.E. Haggard, E.A. Warnemuende, and C. Huang. 2005. Sediment phosphorus dynamics in three tile fed drainage ditches in Northwest Indiana. Agricultural Water Management 71:19-32.

Sparks, D.L. 1999. Soil Physical Chemistry, 2nd ed. Boca Raton, FL: CRC Press.

SPSS, 1999. SPSS Base 10.0 User's Guide. Chicago, IL: SPSS.

Svendsen, L.M., B. Kronvang, P. Kristensen, and P. Graesbol. 1995. Dynamics of phosphorus compounds in a lowland river system: importance of retention and non-point sources, Hydrological Processes 9:119-142.

Taylor, A.W., and H. M. Kunishi. 1971. Phosphate equilibria on stream sediment and soil in a watershed draining an agricultural region. Journal of Agricultural and Food Chemistry 19:827-831.

Thomas, D.L., C.D. Perry, R.O. Evans, E.T. Izuno, K.C. Stone, and J.W. Gilliam. 1995. Agricultural drainage effects on water-quality in southeastern U.S. Journal of Irrigation and Drainage Engineering-ASCE 121:277-282.

Walker, T.W., and A.F.R. Adams. 1958. Studies on soil organic matter: I. Influence of phosphorus content of parent material on accumulations of carbon, nitrogen, sulphur, and organic phosphorus on soils. Soil Science 85:307-318.

Ward, A., D. Mecklenburg, A. Jayakaran, and L. Brown. 2004. Designing two-stage agricultural drainage ditches. Occasional Publication 701P0904. St. Joseph. MI: American Society of Agricultural Engineers.

Yin, C., and Z. Lan. 1995. The nutrient retention by ecotone wetlands and their modification for Baiyandgdian Lake restoration. Water Science and Technology 32:159-167.

Andrew N. Sharpley is a professor in the Department of Crop, Soil, and Environmental Sciences, University of Arkansas, Fayetteville, Arkansas. Tore Krogstad is a professor in the Department of Plant and Environmental Sciences, Norwegian University of Life Sciences, Aas, Norway. Peter J.A. Kleinman is a soil scientist, Francirose Shigaki is a postdoctoral fellow, and Lou S. Saporito is a research support scientist at the Pasture Systems and Watershed Management Research Unit, USDA Agricultural Research Service, University Park, Pennsylvania. Brian Haggard is a professor in the Department of Biological and Agricultural Engineering, University of Arkansas, Fayetteville, Arkansas.
Table 1 Physical and chemical properties of the ditch sediments
collected from the University of Maryland Eastern Shore research farm in
September 2002.

              Particle size analysis  Organic matter
Ditch site    Sand  Silt  Clay        (g [kg.sup.-1])

Forest        95%    2%    3%          2
Mixed         88%    8%    4%          7
Agricultural  47%   36%   17%         77

              CEC                       Mehlich-3 P
Ditch site    (meq 100[g.sup.-1])  pH   (mg [kg.sup.-1])

Forest         3.3                 3.8   29
Mixed          6.8                 3.5   44
Agricultural  13.0                 5.3  107

              Inorganic P                 Oxalate extractable
Ditch site    (mg [kg.sup.-1])  Al (mg [kg.sup.-1])  Fe (mg [kg.sup.-1])

Forest        254                3.1                 32.8
Mixed         518                8.2                 30.8
Agricultural  886               17.4                 54.4

Table 2 Phosphorus sorption properties of fresh (field moisture) and
irradiated ditch sediments from the University of Maryland Eastern Shore
research farm.

              [EPC.sub.0]      Sorption maximum  Binding energy
Ditch site    (mg [L.sup.-1])  (mg [kg.sup.-1])  (L [kg.sup.-1])

Field condition
Forest         0.017           194                 0.326
Mixed          0.032           277                 0.410
Agricultural   0.383           362                 1.217

Irradiated sediments
Forest         0.019           182                 0.215
Mixed          0.036           227                 0.357
Agricultural   0.535           304                 0.551

Change with irradiation
Forest        12%               -6%              -28%
Mixed         13%              -18%              -30%
Agricultural  40%              -16%              -55%

Table 3 Microbial biomass P, calculated as the difference between
inorganic P in triplicate irradiated and nonirradiated sediments, before
and 48 hours after surface runoff from soils receiving poultry litter
was introduced into ditch flow.

                                  Inorganic P
              Before            After             Uptake
Ditch site    (mg [kg.sup.-1])  (mg [kg.sup.-1])  (mg [kg.sup.-1])

Forest        254               272               18
Mixed         518               541               23
Agricultural  886               943               57

                                Microbial P
Ditch site    After (mg [kg.sup.-1])  Percent of P uptake

Forest         4                      22%
Mixed          7                      30%
Agricultural  24                      42%
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Author:Sharpley, A.N.; Krogstad, T.; Kleinman, P.J.A.; Haggard, B.; Shigaki, F.; Saporito, L.S.
Publication:Journal of Soil and Water Conservation
Article Type:Report
Geographic Code:1USA
Date:Jul 1, 2007
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