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a Dep. of Chemistry, Environmental Science and Studies Program, Towson Univ., 8000 York Road, Towson, MD 21252
b Dep. of Environmental Toxicology, Clemson Univ., Clemson, SC 29670
* Corresponding author (racasey{at}towson.edu)
Received for publication September 22, 2000.
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ABSTRACT |
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 TOPNOTES ABSTRACT INTRODUCTIONMETHODSRESULTSDISCUSSIONREFERENCES |
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Abbreviations: ARE, artificial runoff event • LOQ, limit of quantitation
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INTRODUCTION |
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TOPNOTESABSTRACTINTRODUCTIONMETHODSRESULTSDISCUSSIONREFERENCES |
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There are several concerns regarding nutrient leaching into receiving waters. In general, freshwater systems are most sensitive to P inputs while N often limits primary production in estuarine and marine systems. With regard to N leaching, excess NO-3 in surface and shallow ground water can lead to accelerated eutrophication in estuarine as well as freshwater systems (Wetzel, 1983), blooms of noxious algae (Rabalais et al., 1996), and the decline of important seagrasses (Burkholder et al., 1992). Studies sometimes claim that NO-3 leaching is negligible when measured concentrations remain below 10 mg NO3–N L-1 (Cohen et al., 1999; Petrovic, 1990). This concentration is the maximum contaminant level for drinking water (National Academy of Sciences, 1977) but is not necessarily relevant for the protection of ecological structure and function in receiving waters.
In terms of ecological function, some water bodies demonstrate substantial sensitivity to NO-3 contamination even when concentrations are below 10 mg NO3–N L-1. Wetzel (1983) compiled data from a group of experts regarding the N concentrations in eutrophic lakes. Experts reported eutrophic conditions when total N concentrations ranged from 0.4 to 6.1 mg N L-1. The wide range was attributed to different water chemistry in various water bodies. Estuarine systems are generally much more sensitive to excess N than freshwater systems. The National Oceanic and Atmospheric Administration (NOAA) classifies estuaries as having a high eutrophication potential if N concentrations exceed 1 mg N L-1 (National Oceanic and Atmospheric Administration, 1997).
Burkholder et al. (1992) found that NO-3 was highly toxic to the estuarine sea grass Zostera marina L. at concentrations of 2.7 to 4.4 mg NO3–N L-1. This result is particularly important because these sea grasses play a vital role in the health of estuarine systems. Sea grass beds serve as nursery grounds for the larval and juvenile stages of many commercially important finfish and shellfish species, so declines in sea grasses can result in declines of these species as well. Because all surface waters eventually enter estuarine or coastal systems, N management is important throughout the watershed.
Phosphorus has also been shown to contribute to eutrophication, primarily in freshwater systems. While no federal limits on P concentrations in freshwater have been set, the USEPA recommends that total phosphates should not exceed 0.05 mg L-1 and that total P should not exceed 0.1 mg L-1 for water entering lakes or reservoirs (USEPA, 1986). In a review of P and eutrophication, Daniel et al. (1998) stated that concentrations of P between 0.01 and 0.02 mg L-1 were considered levels above which eutrophication was accelerated. In a summary of data for freshwater lakes and reservoirs, Wetzel (1983) compiled values for total P concentrations that correlated with the trophic status of lentic waters. This compilation suggested that total P levels of 0.003 to 0.018 mg P L-1 corresponded to oligotrophic conditions, whereas total P levels of 0.016 to 0.386 mg P L-1 corresponded to eutrophic conditions. These values are similar to the levels of concern expressed by the USEPA (1986).
Much of the research involving riparian zones has focused on the attenuation of NO-3 in shallow ground water (Hill, 1996). Cooper (1990), Simmons et al. (1992), Lowrance (1992), and Jordan et al. (1993) all observed NO-3 attenuation in shallow ground water passing through riparian wetlands. Several studies (Cooper, 1990; Lowrance, 1992) found positive correlations between denitrification potentials and localized NO-3 removal in highly organic wetland soils.
Fewer studies have examined attenuation in wetlands receiving surface inputs of nutrients. Peterjohn and Correll (1984) and Lowrance et al. (1984) both calculated nutrient budgets for riparian areas that included surface and subsurface nutrient inputs. They observed substantial overall nutrient attenuation for both N and P. Ambus and Christensen (1993) investigated NO-3 attenuation and denitrification during storm conditions in a riparian zone receiving tile drainage and runoff from agricultural fields. They observed increases in denitrification rates during storm events and found that the highest rates occurred when the water level was elevated with runoff from the upland soils. Kovacic et al. (2000) quantified NO-3 attenuation in constructed wetlands receiving surface runoff and tile drainage from agricultural fields and observed 37% attenuation of the influent N (mostly in the form of NO-3) over a 3-yr period. They attributed much of their losses to denitrification.
Similarly, many studies of the nutrient retention capacity in wetlands involved systems receiving steady inputs of nutrients from ground water. Fewer studies have examined nutrient attenuation at sites where nutrients inputs occur solely through pulses of storm-generated runoff. A pulse of storm water can rapidly change the electron acceptor status of a soil or ground water by transporting substantial NO-3 loads from terrestrial land uses into systems where more energetic electron acceptors (oxygen) are absent. These intermittent events introduce a temporal component into nutrient attenuation that is not present where the transport of nutrients is more homogeneous. At the site described by Kovacic et al. (2000), although inputs were a combination of storm pulses and base flow, attenuation was observed under both conditions. At the Cheraw State Park site described in this paper, inputs of N and P from the upland land use occurred only during storm events and only through surface transport.
In this study, we quantified the loads of NO-3 and PO3-4 in turfgrass runoff and investigated the efficiency of a riparian wetland for removing nutrients from storm-generated runoff. We then performed controlled amendment experiments to determine whether nutrient attenuation was occurring on the time scale of natural storm events or whether dilution and displacement of pre-event ground water accounted for the observed nutrient attenuation in this wetland. These artificial runoff events (AREs) allowed us to collect a temporal and spatial array of samples within the wetland so that the plume of runoff water moving through the wetland could be traced and the attenuation determined. A companion paper describes the mechanisms by which the nutrient attenuation reported in this paper was achieved (Casey et al., 2001).
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METHODS |
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TOPNOTESABSTRACTINTRODUCTIONMETHODSRESULTSDISCUSSIONREFERENCES |
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At the upstream side of the wetland, Transect 1 contained a Johnston series organic soil (coarse-loamy, siliceous, active, acid, thermic Cumulic Humaquept) alongside a Dorovan series organic soil (dysic, thermic Typic Haplosaprist). The rest of the wetland was a Dorovan series organic soil except for a small area of Pamlico series organic soil (sandy, siliceous, dysic, thermic Terric Haplosaprist) near Sampling Site 4-3. At the upstream side of the wetland, the depth of the organic horizon averaged 0.5 m. The depth of this layer increased to 3 m toward the downstream end of the wetland. The organic matter content of this layer averaged 18%.
Storm Event Sampling
A U.S. Geological Survey stream gauging station equipped with a Sigma automatic water sampler (American Sigma, Medina, NY) was installed upstream of the wetland and adjacent to the drainage ditch that carried runoff water from the turfgrass to the wetland (Fig. 1). This station was identified as the inlet. A rectangular notch weir was installed in the ditch to facilitate measurement of the discharge that occurred during storms. A Sutron (Sterling, VA) 8200A data recorder collected stream stage values at 15-min intervals regardless of storm conditions. Stages were measured using a float located in a stilling well. Sampling initiated after the stage reached a predetermined trigger level and proceeded at 10- to 30-min intervals until 24 samples were obtained. Sampling for reported events always included the rising and falling limb of the hydrograph, although discharges did not always returned to base flow (or zero discharge for the inlet) by the end of sample acquisition. Event volume was determined for the time period during which sampling occurred using the inflow at the inlet. Event volumes were calculated by integrating the instantaneous discharges over time for the duration of sampling.
A second station (identified as the outlet) was installed downstream of the riparian wetland to characterize water discharging from the system. Upwelling from the wetland resulted in a first-order stream that flowed approximately 20 m before entering Lake Juniper. Because the bottomland area was flat and the stream was only 10 to 15 cm deep, it was not possible to install a control device such as a weir to facilitate discharge measurement. Instead, a stage–discharge relationship was developed using the stream channel itself. A staff plate and gauging device were installed at the sampling site and stream discharge was measured with a digital flow meter over a range of flow conditions. Measured discharges and the corresponding stream gauge heights were used to generate a stage–discharge relationship. Interpolation between the measured points allowed for quantification of stream discharge. Stages were recorded digitally as described for the inlet.
Storms were not sampled based on criteria of size, duration, or season. Sampling of a storm was based on the status of the equipment and the availability of personnel who retrieved the samples from the field. This strategy resulted in a wide range of storm characteristics for the sampled events. Figure 2 shows the distribution of sampled events in the context of the range of events throughout each water year (1 October–30 September). Loads of NO-3 and PO3-4 were calculated using the measured concentrations and the instantaneous discharge at the time of sampling. When concentrations were below the limit of quantitation (LOQ), a range was reported for the load, indicating that the concentration could be between zero and the LOQ. Attenuation, determined as the percent loss between the inlet and outlet, was also reported as a range for these cases. Volume-weighted nutrient concentrations were calculated by dividing the total load of nutrient measured during an event by the total event volume measured at the inlet.
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The ground water samplers consisted of 30-cm lengths of 5-cm-diameter PVC pipe with 0.25-mm slotted screen. The bottom of the sampler was a 5-cm PVC cap and the top was a rubber stopper with two 1-cm holes for tubing. Two pieces of vinyl tubing were inserted into the rubber stopper and were kept closed above the ground surface with plastic hose clamps. One length of tubing extended to the bottom of the sampler and was used to withdraw water samples. The other length of tubing only extended to the top of the sampler and was used for pressure relief to prevent the formation of a vacuum inside the sampler as water was withdrawn. The entire length of each sampler was screened, so water samples represented a depth integration of 0.3 m. Samplers were installed with a hand auger. A 2.5-cm sand pack was installed around the sampler to prevent clogging and to increase the effective radius from which the samplers were able to draw ground water. Excavated soil was returned to the auger hole above the samplers in an attempt to preserve the local soil chemistry. Bentonite clay pellets were placed in the upper 15 cm of the auger hole to prevent vertical channeling of surface water.
Four transects of ground water samplers were installed within the wetland, perpendicular to the direction of ground water flow (Fig. 1). Each transect contained three nest sites with two or three depths at each site. All nest sites had samplers extending from 0.3 to 0.6 m and 1.2 to 1.5 m below ground surface. Several sites at Transects 3 and 4 also had samplers at 2.1 to 2.4 m below ground surface because of the greater depth of the Dorovan series soil at these transects. Because the samplers were deployed in a saturated soil, each was emptied and placed under a 0.8-kPa vacuum before sampling. Samples were obtained after 15 min of recharge. This ensured that samples were representative of current ground water conditions.
Artificial Runoff Events
We conducted AREs designed to mimic the intermittent loading of nutrients experienced by the wetland at this site. Using AREs provided the opportunity to conduct extensive sampling of surface and ground water during an event. We used the network of ground water samplers to determine the spatial and temporal trends of nutrient attenuation within the wetland. Previous experience demonstrated that surface water entering the wetland infiltrated rapidly into the subsurface of the sandy shelf area and Transect 1 (Fig. 1). We expected amended surface water to enter the subsurface in this area and then continue through the wetland subsurface. The flow rates chosen for the AREs resulted in smaller event volumes than in many of the natural storm events. This was done intentionally to examine the fate of infiltrated runoff without the confounding effect of substantial overland flow. Most of the natural events were sampled over 12 h whereas the amendment period for AREs 1 and 2 was only 6 h, contributing to the apparent difference in event sizes between the natural and artificial runoff events.
The AREs were conducted by pumping water from an irrigation system into the existing ditch that directed runoff from natural storm events toward the upstream side of the wetland (Fig. 1). This water was amended with NO-3, PO3-4, and Br-, which mixed with the pumped water in the ditch and then flowed to the sandy shelf. The irrigation system pumped its water from a lake (Lake Juniper) with water chemistry similar to the ground water within the wetland. The water was pumped directly into the ditch that normally carried natural storm runoff. In contrast to natural storm events, no precipitation fell directly onto the wetland during these events; all of the water entered through surface flow.
Sodium nitrate, monobasic sodium phosphate, dibasic sodium phosphate, and sodium bromide were weighed and dissolved to near saturation before the amendment commenced. The salt concentrates were diluted up to 200 L in a carboy for delivery into the ditch by a peristaltic pump at a constant rate. The results of runoff monitoring indicated that turfgrass runoff exhibited relatively constant concentrations of NO-3 and PO3-4 during the course of individual storms and did not demonstrate a significant first-flush effect.
A 5-m length of ditch served as a zone of mixing downstream of the amendment site, upstream of the first surface water sampling station (the inlet). Water samples were collected at the inlet at 15-min intervals during and immediately following the amendment. Discharge was also measured at 15-min intervals at this station. Samples were obtained at the outlet every 30 min and discharge was measured every 15 min for the duration of each experiment. As in the sampling of natural runoff events, these stations provided a measure of the nutrient composition of runoff water entering and leaving the wetland, respectively.
During the AREs, ground water samples were obtained at 2-h intervals during daylight on the day of amendment and at 4-h intervals during daylight on the day after amendment. Artificial Runoff Event 3 also included two post-amendment sampling times during the third day. This sampling scheme allowed us to develop a detailed spatial and temporal characterization of the runoff plume within the wetland.
The composition and duration of runoff varied for the three AREs (Table 1). The nutrient loads used in these experiments were based on data from natural runoff events at this site. For the first experiment (ARE 1), we chose NO-3 and PO3-4 loads that were similar to the highest levels found in the natural events sampled in 1996 and 1997. In order to investigate the upper limit of NO-3 attenuation in the wetland, ARE 2 and 3 introduced approximately twice as much NO-3 as was measured in any of the natural storms (Table 1).
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Sample Analysis
Sample pH was determined in the field immediately following acquisition using a pH electrode. A subsample was then filtered through a 0.2-µm PTFE syringe filter (Gelman Laboratory, Ann Arbor, MI) for analysis by ion chromatography at the laboratory. The remaining sample was preserved with several drops of concentrated sulfuric acid to bring the pH to 2.0 or lower (American Public Health Association, 1989).
Nitrate, nitrite, PO3-4, and Br- were quantified using an ion chromatograph (Millipore-Waters, Milford, MA) consisting of a WISP 710B autosampler, 510 pump, 431 conductivity detector, and an IC-PAK Anion HR column. The limits of quantitation for these ions were: 0.1 mg NO3–N L-1, 0.15 mg NO2–N L-1, 0.2 mg PO4–P L-1, and 0.5 mg Br L-1.
For AREs 1 and 2, ammonium N and total P were quantified for every fourth sample collected at the inlet and every other sample from the outlet (60 min intervals), all initial ground water samples, all ground water samples from the last sampling time on Day 1 of the ARE, and all ground water samples from the final sampling time. Ammonium was quantified using the spectrophotometric phenate method (American Public Health Association, 1989) with a limit of quantitation of 10 µg NH3–N L-1, and total P was quantified using the ascorbic acid–molybdate method (American Public Health Association, 1989) following digestion of the sample with 5 mL concentrated nitric acid and 1 mL concentrated sulfuric acid. For ARE 1, 10 mL of sample were digested, resulting in a limit of quantitation of 50 µg P L-1. We used 40 mL for the digestion in ARE 2, which lowered the limit of quantitation to 12 µg P L-1. In both cases, this method provided a lower detection limit for P than the ion chromatographic analysis of PO3-4. This allowed for a more sensitive determination of P transport trends during simulated runoff conditions. Every other sample (60-min intervals) from the outlet was also analyzed for total Kjehldal N for AREs 1 and 2 (American Public Health Association, 1989).
The concurrent amendment of NO-3, PO3-4, and Br- allowed us to determine nutrient attenuation by comparing NO-3 and PO3-4 dynamics to the behavior of the Br- tracer. Bromide acts as a useful conservative tracer in soil (Smith and Davis, 1974). The degree of nutrient attenuation was quantified using ratios of the nutrients to the Br- tracer (Trudell et al., 1986). The percentage loss of nutrients to attenuation was determined as follows:
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RESULTS |
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TOPNOTESABSTRACTINTRODUCTIONMETHODSRESULTSDISCUSSIONREFERENCES |
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Nitrate and PO3-4 were detected in turfgrass runoff at the inlet in all of the sampled natural storm events (Tables 2 and 3). Of the 12 storms, nitrite (NO-2) was present in the 7 Oct. 1996 and 25 Sept. 1997 events with loads at the inlet of 189 and 102 g NO2–N, respectively. Nitrite was not detected in any other events.
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Both NO-3 and PO3-4 loads in storm water were attenuated after passing through the wetland at this site. Nitrate attenuation averaged 80% (including uncertainties for samples with concentrations less than the limit of quantitation) in the 11 storms sampled for attenuation. Neither NO-2 nor PO3-4 was ever detected at the downstream sampling station. Including uncertainties for sample concentrations less than the LOQ, the average PO3-4 attenuation was 74%. In general, the large range of attenuation for PO3-4 in several of the storms was due to the small loads measured at the inlet. In these cases, the uncertainty associated with sample concentrations below the LOQ translated to a large percentage of the load that entered the wetland. In the storms that transported the largest loads of PO4–P, attenuation was highest and had a tighter range. These storms probably give the best indication of the true PO3-4 retention capabilities of this wetland.
The storms with the largest event volumes generally had the largest exports of NO-3 from the wetland (Table 2). This is probably due to the fact that larger storm volumes result in shorter retention times. In addition, observations during storm events indicated that high flow rates exceeded the infiltration capacity of the sandy shelf area resulting in hydraulic overloading and overland flow through portions of the wetland. In storms where direct observations were available, overland flow did not extend continuously from the inlet to the outlet; thus, storm water bypassed only a fraction of the wetland soil. These soils were probably the site of substantial nutrient attenuation, as described in the companion paper (Casey et al., 2001).
Artificial Runoff Events
Preliminary observations from the ground water samplers indicated that NO-3 and PO3-4 were not present in ground water in the interim between storm events (data not shown). Ground water samples were obtained on seven occasions between July 1997 and February 1998 and PO3-4 was never detected (detection limit 0.2 mg PO4–P L-1) in the wetland ground water. Nitrate was detected repeatedly at one sampler in Transect 1, possibly as a result of infiltrated storm water seeping into the wetland from the sandy shelf after surface flow had ceased. Nitrate was not detected at the other transects during the time period between storms. These observations confirmed that NO-3 and PO3-4 from infiltrated storm runoff were not being stored in ground water between events. These observations also suggested that the wetland was receiving nutrient inputs predominantly in pulses during storm events and not through continuous ground water transport, despite discharge measurements indicating a net discharge of ground water within the system.
In all three AREs, runoff water entering the wetland infiltrated into the subsurface and dispersed both vertically and laterally. No continuous flow of surface water reached the outlet, but the flow rate did exceed the infiltration capacity of the sandy shelf area resulting in a flow of surface water that reached the area between Monitoring Transects 1 and 2 before seeping into the subsurface. The Br- plume was detected at all of the sampling sites of each ground water monitoring transect (Fig. 3) and at the outlet (Fig. 4), confirming the hydraulic connection between the upstream and downstream sampling stations. Detecting Br- at these sites allowed the observed losses of NO-3 and PO3-4 to be quantified as attenuation because the determination of attenuation requires the observation of Br-. In addition, the tracer confirmed that the direction of ground water flow was from the inlet to the outlet because Br- was seen to move from the site of application sequentially from Transect 1 to Transect 4 and the outlet.
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The vertical plume distribution was also uneven at each of the ground water monitoring transects. The ARE water remained in the shallow subsurface and did not exhibit extensive vertical dispersion. The highest concentrations of Br- intercepted the uppermost ground water samplers throughout the wetland. Bromide concentrations in the deeper samplers (>1.2 m below ground surface) were lower than their shallow counterparts at almost every sampling time and location for all three AREs (Fig. 3 and 5).
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Extensive attenuation of PO3-4 occurred in both AREs in which PO3-4 was applied (AREs 1 and 2). Phosphate was only detected in one ground water sample at Transect 1 in each experiment. Surface water that had not yet infiltrated in the wetland had PO3-4 attenuation of 53 to 85% in ARE 1 and 30 to 72% in ARE 2 (data not shown). In all other samples from AREs 1 and 2, PO3-4 was not detected. Phosphate was never detected past Transect 2 in either of the AREs, indicating extensive retention of the applied PO3-4 in the first 30 m of the wetland. Total P concentrations in surface water at the outlet remained at or below the analytical P detection limit, confirming the retention of PO3-4 in the wetland (Table 5). This also indicates that the plume of applied inorganic P was not transformed into a plume of organic P during the time scale of this experiment. Because of the low LOQ for the total P determination (0.05 mg P L-1 for ARE 1; 0.01 mg P L-1 for ARE 2), dilution alone, as quantified by change in Br- concentration, cannot account for the loss of P. If only dilution of the ARE water were occurring, we should have observed maximum total P concentrations near 1.0 mg P L-1 at the outlet in association with the maximum Br- concentrations. Total P measurements were at least 20-fold less than this amount based on the LOQ.
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In terms of recovery of the applied tracer, the mass balances for the three AREs were moderate (Table 6). The storage term in Table 6 was determined using the amount of each substance remaining in the wetland at the end of each ARE (Table 7). Given the wetland dimensions of 30 by 200 m, a sampling depth of 0.3 m for each level of sampler, and assuming a porosity of 50% in the organic wetland soils, estimates of the mass remaining in the subsurface were calculated. The storage term probably underestimates the amount of storage because there was no data for the 0.0- to 0.3- and 0.3- to 1.2-m-deep portions of the wetland. Little storage was present in the 1.2- to 1.5-m-deep portion of the soil profile and none occurred in the 2.1- to 2.4-m-deep areas. The storage plus output terms of the mass balance accounted for 39 to 57% of the applied Br- in the AREs. It is likely that a large fraction of the missing Br- was present in the 0.0- to 0.3-m-deep portions of the wetland. Bromide was detected in several small surface pools in the vicinity of the sampling transects; however, because a volume could not be estimated for these pools, their contribution could not be included in the storage term. In contrast to the Br-, none of the applied PO3-4 and little of the applied NO-3 remained in storage in the ground water at the conclusion of the experiment. Of the NO-3 that was present at the conclusion of each experiment, almost all of it was present at Transect 1 where small amounts of dosing material were still entering the wetland subsurface from the sandy shelf. Thus, the NO-3 that was present at the end of the experiment had yet to be transported through the wetland.
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DISCUSSION |
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TOPNOTESABSTRACTINTRODUCTIONMETHODSRESULTSDISCUSSIONREFERENCES |
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In the AREs, the presence of Br- at the outlet confirmed the hydraulic connection between the two surface water sampling stations. This also indicated that the attenuation inferred by NO-3 disappearance between the inlet and outlet during natural runoff events did not result from nutrient-rich surface water seeping into deeper aquifers that were unrelated to the water discharging from the wetland. Instead, the Br- distribution indicated that surface water entered the wetland, dispersed through the shallow subsurface, and left the wetland through the first-order stream originating at Transect 4.
The observed dilution of Br- probably resulted from mixing between amended runoff water and pre-event wetland ground water. Waddington et al. (1993) observed similar mixing in a bottomland wetland. They found that much of the mixing occurred in preferential flow macropores. The presence of preferential flow macropores in this wetland would explain the observed dilution as well as the rapid subsurface transport of tracer along the 200-m distance of the wetland to the outlet. Breakthrough of Br- occurred in less than 10 h in all of the AREs as shown in Fig. 4. Preferential flow macropores were observed during installation of the ground water monitoring devices in this wetland and may be contributing substantially to its hydrology.
It appeared that little of the amended water from the inlet left the wetland in ground water flow at Transect 4. Bromide concentrations were much higher in the surface water at Transect 4 (as measured at the outlet) than in the ground water (Table 4). This is consistent with the observation that Transect 4 was a site of ground water discharge, as evidenced by the first-order stream that originated at this area. Visual observation of surface flow as well as hydraulic head measurements in subsequently installed piezometers (data not shown) showed that upwelling at Transect 4 occurred continually, both during and between storm events. This upwelling was probably a result of the hydraulic proximity of the adjacent lake and resulted in the plume of amended water returning to the surface at, or before, Transect 4.
These experiments demonstrated that nutrient attenuation in this wetland occurred on the time scale of natural storm events. In contrast to reports dealing with steady ground water inputs of NO-3 over long periods (Simmons et al., 1992; Lowrance, 1992), we observed the initiation and completion of nutrient attenuation within a 24- to 48-h period. This indicated that the mechanisms responsible for nutrient attenuation persisted between storm events and could be rapidly initiated upon the introduction of nutrients in runoff.
We observed a rapid increase in NO-3 attenuation from a low value at the time of breakthrough to a plateau level near 100%. Given that this wetland was composed of saturated soils with high organic matter content, microbial denitrification is probably one of the dominant mechanisms for NO-3 attenuation. Smith and Tiedje (1979) noted that when denitrification was initiated following an aerobic period, there were several phases in the observed activity. Initially, only existing enzymes were available for NO-3 reduction in a period that lasted 1 to 3 h in laboratory experiments. Following this phase, de novo enzyme synthesis occurred and the rate of denitrification increased. Parsons et al. (1991) and Schipper et al. (1994) also observed that denitrification activity persisted in soils after periods that were not conducive to denitrification. Thus, denitrification in this riparian wetland could well account for the observed magnitude and dynamics of NO-3 attenuation.
Phosphate attenuation occurred immediately after the initiation of the ARE amendments. Several possible mechanisms for PO3-4 attenuation in wetlands have been proposed, including sorption to soil mineral matter (Richardson, 1985; Gale et al., 1994), uptake by plants (Adler et al., 1996), and uptake by microbes. Any of these mechanisms could account for the attenuation observed in this experiment. However, the amount of PO3-4 removed over a very short time period suggests that sorption was the dominant mechanism over the short term. Sorption was quantified in the companion paper (Casey et al., 2001). However, as Richardson (1985) noted, sorption is a mechanism that can be saturated. Although this was not directly measured in our experiment, we speculate that the magnitude of PO3-4 retention in this wetland may currently rely in part on plants and microbes removing sorbed PO3-4 from mineral surfaces in order to renovate those binding sites for subsequent storm events. In that case, the wetland may serve as a P sink until the vegetative biomass reaches a climax state, after which P exports would be expected.
In conclusion, storm event runoff at this site contained NO-3 and PO3-4, which in some events reached levels that could result in negative effects on receiving waters. Both nutrients were retained by a natural riparian wetland preventing their migration into adjacent water bodies. The AREs demonstrated that the nutrient attenuation measured during natural runoff events could be attributed to processes occurring in the riparian wetland. Nutrient attenuation of ARE water occurred on the time scale of natural runoff events (6–48 h) and resulted in the removal of both NO-3 and PO3-4. The conservative Br- tracer used in the AREs indicated that substantial mixing of runoff water and pre-event ground water occurred at the wetland subsurface. The tracer also indicated that the majority of runoff water remained in the shallow subsurface and did not exhibit dispersion beyond 1.2 m below ground surface. The experimental approach of using AREs allowed us to simulate runoff conditions and successfully measure attenuation under field conditions that were close to those occurring in natural events. We were able to determine that nutrient attenuation occurred in the wetland without the necessity of extrapolating this function solely from soil properties measured in the laboratory. A companion paper (Casey et al., 2001) provides insights into the mechanisms that contributed to attenuation at this site.
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ACKNOWLEDGMENTS |
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NOTES |
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TOPNOTESABSTRACTINTRODUCTIONMETHODSRESULTSDISCUSSIONREFERENCES |
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REFERENCES |
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TOPNOTESABSTRACTINTRODUCTIONMETHODSRESULTSDISCUSSIONREFERENCES |
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This article has been cited by other articles:
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  G. Vellidis, R. Lowrance, P. Gay, and R. K. Hubbard Nutrient Transport in a Restored Riparian Wetland J. Environ. Qual., March 1, 2003; 32(2): 711 - 726. [Abstract] [Full Text] [PDF]
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R. E. Casey, M. D. Taylor, and S. J. Klaine Mechanisms of Nutrient Attenuation in a Subsurface Flow Riparian Wetland J. Environ. Qual., September 1, 2001; 30(5): 1732 - 1737. [Abstract] [Full Text] [PDF]
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