Estimating nitrate load reductions from placing constructed wetlands in a HUC-12 watershed using LiDAR data

Ecological Engineering 56 (2013) 69–78

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Estimating nitrate load reductions from placing constructed wetlands in a HUC-12 watershed using LiDAR data M.D. Tomer a,∗ , W.G. Crumpton b , R.L. Bingner c , J.A. Kostel d , D.E. James a a

USDA/ARS, National Laboratory for Agriculture and the Environment, Ames, IA, United States Iowa State University, Dept of Ecology, Evolution & Organismal Biology, Ames, IA, United States c USDA/ARS, National Sedimentation Laboratory, Oxford, MS, United States d The Wetlands Initiative, Chicago, IL, United States b

a r t i c l e

i n f o

Article history: Received 13 October 2011 Received in revised form 26 March 2012 Accepted 29 April 2012 Available online 9 June 2012 Keywords: Light detection and ranging (LiDAR) Nutrient-removal wetlands Nitrate Precision conservation AnnAGNPS

a b s t r a c t Mitigating Gulf of Mexico hypoxia will require practices to reduce nitrate losses from tile drainage throughout the upper Mississippi River basin. Wetlands are a key practice to accomplish this via denitrification, but locations of feasible wetland sites will need to be determined on a watershed specific basis. This study’s objective was to demonstrate that LiDAR topographic data can be used to identify feasible wetland locations in a 6500 ha watershed in northern Illinois, and then estimate the impact on watershed nitrate loads from wetlands hypothetically constructed at those locations. The evaluation resulted in the identification of eleven sites where wetlands could intercept tile drainage from 30% of the watershed. The USDA AnnAGNPS model was used to estimate nitrate loads delivered to candidate wetlands during a 30-year period of simulated climate. The model results were consistent with discharge and nitrate loading regimes in the region. Nitrate reduction by the wetlands was estimated using a published regression model. The wetlands could reduce average watershed nitrate loads by 11–13%, a significant reduction from treating just 30% of the watershed with a single practice. Results showed a wide variation in N-removal performance among wetlands, due to varying contributing-to-wetland area ratios, and varying land uses that affected discharge and N loads among locations. A large wetland installed at the watershed outlet could achieve an estimated N-load reduction of 35% under relaxed siting criteria, but would require the added expense of roadway reconstruction. These conclusions are relevant for planning watershed conservation efforts and establishment of nutrient trading schemes. Published by Elsevier B.V.

1. Introduction Decreasing nitrate loads discharged by tributary rivers in the Upper Mississippi River (UMR) basin is an important goal for mitigating Gulf of Mexico hypoxia (USEPA, 2008). The targeted reduction of 45% (Dale et al., 2010; Justic´ et al., 2007) will be difficult to achieve without broad measures that include optimized nutrient management within agricultural fields, and approaches to remove nitrate in waters draining from those fields. Artificially (tile) drained areas of the Midwest are losing disproportionate amounts of nitrate-N (NO3 -N) and any effective nitrate-reduction strategy will need to specifically address tile-drained land (Royer et al., 2006). Wetlands are one of the most effective practices that can remove nitrate from surface waters that receive tile drainage (Crumpton et al., 2006; Dinnes et al., 2002; Kovacic et al., 2000), through the process of denitrification. Therefore, the installation of

∗ Corresponding author. E-mail address: [email protected] (M.D. Tomer). 0925-8574/$ – see front matter. Published by Elsevier B.V. http://dx.doi.org/10.1016/j.ecoleng.2012.04.040

wetlands to receive and treat tile drainage is likely to be a critical component of any successful nitrate reduction strategy for the UMR basin (Crumpton et al., 2006, 2008; Mitsch et al., 2005). But what is the possible extent of wetlands in tile-drained watersheds, and what amount of nitrate might be removed if they were installed? While it is clear that nutrient removal in wetlands is most effective if they are placed to receive significant volumes of tile drainage and sized appropriately (Crumpton, 2001), exact answers will certainly be unique to each watershed. It is important to understand the potential of this cornerstone practice as part of a comprehensive strategy to meet the nitrate reduction goal in any given watershed. In tile drained watersheds, only certain locations are feasible for installation of nutrient-removal wetlands that will receive tile drainage from relatively large agricultural areas that discharge tile drainage. These wetlands must also be sized to provide average hydraulic residence times on the order of several days, rather than hours or weeks, to the wetland area’s effectiveness for nitrate removal. Wetlands additionally must be placed and designed to avoid impeding subsurface drainage from significant areas of tile drainage, thereby causing them to be removed from crop

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production. Finally, impacts on infrastructure (i.e., roads, farmsteads, etc.) would probably be unacceptable to local stakeholders. Therefore, in any watershed dominated by tile-drained lands, identifying the feasibility of wetlands and their potential for nitrate removal, considering these constraints, will help to develop alternatives for watershed-specific nutrient reduction strategies. This paper provides a case study illustrating a method to map the most feasible locations for nutrient-removal wetlands under a specific set of criteria that are based on topographic data, and estimate their potential contribution toward nitrate reduction in a 6500 ha watershed in the upper Mississippi River basin. Topographic information is critical to identify locations for many conservation practices such as terraces (Arabi et al., 2008), grassed waterways (Pike et al., 2009), riparian buffers (Dosskey et al., 2011), and other practices that can reduce soil erosion and/or trap sediment and nutrients. Indeed, topographic data are required to implement any USDA conservation practice that must be site specifically designed, according to slope length and steepness (Renard et al., 1997). However, acquisition of high-resolution topographic survey data is expensive, and therefore data have usually been collected only as needed, on an individual site basis. But technology has changed that limitation, and airborne laser altimetry can now provide detailed topographic data across large areas. These data, called light detection and ranging (LiDAR) data, are becoming increasingly available. The utilities of LiDAR data are many and include mapping of floodplains and geological features (Jones et al., 2007), conducting archeological surveys (Crow et al., 2007), and mapping vegetation features, including tree canopy heights (Hawbaker et al., 2010) and wetland areas beneath forest cover (Lang and McCarty, 2009). Clearly, LiDAR data offer opportunities to map agricultural fields and precisely apply conservation practices across watersheds in order to meet specific environmental goals. Terrain attributes derived from topographic data can help identify/prioritize sites for nutrient removal wetlands and riparian buffers in watersheds (Tomer et al., 2003), but LiDAR offers a new resource to use site-specific analyses for managing agricultural water quality at the watershed scale. The purpose of this paper is, first, to demonstrate how data from a watershed-scale LiDAR survey can be used to locate sites for nutrient-removal wetlands that meet specific criteria. Second, the potential contribution of these wetlands, if installed, toward reducing the nitrate-N load discharged from the watershed is evaluated through a modeling exercise. Results will have implications that include prioritizing wetland sites for incentive payments and evaluating the feasibility of including these wetland sites in nutrient trading markets.

2. Methods 2.1. Study location Big Bureau Creek (BBC) is in north-central Illinois and is tributary to the Illinois River. The BBC drains a 1280 km2 watershed comprised of glacial deposits of Wisconsinan-age (12,000–22,000 yr b.p.). This study was conducted in Lime Creek (Fig. 1), a northern tributary to Big Bureau Creek with a watershed drainage area of 6500 ha. The watershed is dominated by thick (>100 m) glacial deposits; it is bounded to the north by a terminal moraine, with a glacial-fluvial plain covering the southern half of the watershed. The till is typically capped with a loess mantle averaging about 1.5 m thick. The area was settled by Europeans during the 1830s, who converted the native prairies and wet meadows to agricultural use within a few decades. Improved drainage was necessary to achieve this transition and a system of ditches and subsurface (tile) drains were installed to make agricultural production

feasible. Today, the watershed is drained by a network of straightly dug drainage ditches, and the dominant land use is for corn and soybean production. 2.2. Acquisition and processing of survey data LiDAR data were obtained for the Lime Creek watershed by aircraft during December 2008, at a time with no snow cover, to best enable representation of a ‘bare-ground’ surface. Final return data from laser pulses were converted to a digital terrain model of elevation points, and then processed to obtain a 1-m grid digital elevation model (DEM) of the land surface (Fig. 1). This DEM was subjected to hydrologic modeling of overland flow and modified (as described below) to enforce flow pathway convergence into drainage ditches and downstream. The modeling was conducted using ArcGIS version 9.2 software, Arc Hydro tools (ESRI, 2009). The first step was a “pit-filling” operation to fill depressions and estimate where and in what direction filled depressions would overflow. A surface of fill depths was generated by subtracting the original DEM from the filled DEM. This information identified natural depressions that are common on glaciated terrain, but it also showed artificial impoundments up-gradient of roadways, because LiDAR data cannot represent how flow occurs through culverts and beneath bridges. Rather, in the original LiDAR data, hydrologic routing errantly impounded flows at bridges and routed the flows over road surfaces. This requires the LiDAR data to be manipulated to ensure overland flows are correctly routed through roadway infrastructures. Artificial impoundments formed by roadways were removed by the following procedure. First, intersections between roadway centerlines and ditches were digitized as points, based on interpretation of original, unfilled DEM imagery. The road surface within a set radius of these points was then decreased (or ‘burned’) to 1 m below the channel elevation. To accomplish this, a ‘burn’ depth map layer was established that included nil values for all of the watershed except those grid cells within the set radius of the road-ditch intersections. Within this set radius, the maximum fill depth (obtained using the ‘filled’ DEM as described above) found within that radius of the digitized intersection points, plus 1 m, was assigned to all grid cells within the search radius. The search radius distance was chosen according to the type (width) of the roadway; a 15 m radius sufficed for most roads but a 20 m radius was required along a wider state highway that crossed the watershed. The ‘burn-depth’ map layer was then subtracted from the original DEM grid-cells, lowering the elevation within the search radius of road-channel intersections, but leaving the rest of the elevation data throughout the watershed unaltered. An example of one ditchroadway intersection is shown in Fig. 2, illustrating the original DEM and the result of the ‘burn’ operation to remove the false impoundment. The fill operation and hydrologic flow modeling was then repeated on the ‘burned’ DEM. This procedure successfully maintained the simulated routing of accumulated overland flows within drainage ditches. The burn process was repeated for several locations that were not along drainage ditches, in order to accurately include some areas as being within (or out of) the watershed. This was necessary where roadways were built above (about 1 m) the flat terrain of Lime Creek’s glacial-fluvial plain to help accommodate farm traffic. 2.3. Identifying possible nutrient removal wetlands Once actual surface flow pathways within the ditch channels were accurately routed, a search for candidate wetland sites was undertaken. Potential wetland sites occur along drainage ditches that were dug into alluvial sloughs (see Fig. 3). Prior to

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Fig. 1. Shaded relief map of Lime Creek in northern Illinois, based on LiDAR topographic data. Roads and channels with >100 ha contributing area are also shown.

settlement, these sloughs were weakly or non-channelized, occupied by wet meadows, and provided slow conveyance of surface flows toward Bureau Creek. The sloughs are currently cropped or in pasture, but wet conditions frequently restrict planting and harvesting schedules, reducing their farm-ability but making them potentially suited for wetland installation. The relatively suitability of candidate wetlands can only be evaluated according to consistent criteria. Some choices in initially setting the criteria are arbitrary, and depend on the type of design for the wetland. First, the smallest contributing area that would be considered for wetland installation had to be defined. Wetlands with small contributing areas may be ineffective in decreasing nitrate loads at the watershed scale (Crumpton, 2001). A minimum of 100 ha was selected as the minimum contributing area

considered for wetland assessment for this study. This threshold is arbitrary, and smaller than the minimum 200 ha criteria used to site wetland in the Iowa Conservation Reserve Enhancement Program (Tomer et al., 2003). Evaluation of aerial photographic images (National Agriculture Imagery Program, 2010) indicated that channelized ditches generally initiated near this area threshold. Later, a field review confirmed that flow paths with <100 ha contributing area were dominantly ephemeral and non-channelized (i.e., grassed waterways). Therefore, virtually the entire network of ditches draining Lime Creek was screened for wetland suitability, as well as a large subsurface drainage main in a southeast sub-basin. Candidate wetland sites were constrained by a conservative set of criteria that could initially be accepted by landowners whose livelihood depends on continued artificial drainage of

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Fig. 2. Left panel: shaded relief image of a roadway-channel intersection. Modeling hydrologic flows of this surface requires the area upstream (north) of the roadway to be ‘filled’ and then flow over the road, which typically spills into the down-gradient field and misrepresents the actual hydrology. Right panel: the road-channel intersection has been ‘burned’ as described in the text, which allows digital routing of hydrologic flows to stay in the channelized ditch.

cropland. The basic design of the wetland was chosen to be a low impoundment that would flood a length of drainage ditch and some area of adjacent low-lying land. Several basic approaches to wetland construction are possible, including stream diversions with off-channel treatment (e.g., Moreno-Mateos et al., 2010), but in small watersheds like Lime Creek (Fig. 3), off-channel features like oxbows and abandoned channels are not present, therefore off-channel wetland construction would require excavation. The alternative of impounding the ditch takes advantage of the natural landscape, providing esthetic value and minimizing earthmoving requirements. With the detail of LiDAR, the impoundment depth could be specified as the height of ditch bank plus some constant (selected to be 0.9 m) to provide for a flooded wetland area. The 0.9 m depth across most of the wetland enables emergent vegetation to readily establish, which stimulates carbon fixation that is required for denitrification. A ‘buffer’ around the wetland, where the surface elevation is within 1.5 m of the wetland pool elevation, was specified to account for the need to ensure free drainage of the upslope area. Tiles are typically installed at 1.2 m depth below the surface. The criteria of a maximum wetland depth of 0.9 m and wetland

buffer within 1.5 m of the wetland’s pool elevation were modified from screening criteria used in Iowa’s Conservation Reserve Enhancement Program (CREP), discussed by Tomer et al. (2003). The wetland buffer criterion also ensures that wetland installation would present virtually no risk of exacerbating local flooding even during extreme events. Finally, wetlands would only be recommended where they would not impact infrastructure within the watershed; i.e., drainage beneath and along roads in the watershed had to remain unimpeded by wetlands at the maximum buffer elevation. To identify suitable wetland sites, an impoundment was digitized just up-gradient of each road-channel intersection that was 2.4 m (i.e., 0.9 m wetland pool plus 1.5 m buffer) greater in height than the ditch-bank elevation. The ditch bank elevation was estimated by adding the focal range (range in elevation within 20 m of the channel) to the channel elevation, which was estimated by the focal minimum elevation. If the focal minimum (i.e., estimated channel elevation) at the next up-gradient road crossing was greater than this “buffer impoundment” elevation, then that location was tallied as a candidate wetland. Candidate sites that passed this check then had a lower impoundment simulated that was the

Fig. 3. A typical drainage ditch in the Lime Creek watershed. Candidate wetland sites occurred where ditches could be impounded while impeding subsurface discharge (through drains installed throughout the watershed) from only small areas.

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focal minimum plus the focal range plus 0.9 m. This resulted in two impoundments, the smaller representing the potential wetland pool and the larger representing the wetland plus the surrounding buffer. Most of these buffer areas would likely be removed from row-crop production, requiring purchase of a land-use easement from the landowner. A final check was made using rectified aerial photographs to ensure the wetlands and buffers did not overlap any farmsteads or other infrastructure. The potential wetland sites identified by this process were reviewed in the field to evaluate the approach. While the site selection criteria described above were conservative, they could be adjusted based on experience in implementing watershed-scale, wetland placement strategies, and broadened acceptance by landowners. After identifying the candidate sites, areas of wetlands, buffers, and contributing areas were tabulated. Ideally, wetland sites should be located to intercept drainage from relatively large areas, and be sized between 0.5% and 2.0% of the contributing area (based on Iowa CREP criteria, see Tomer et al., 2003). A preliminary ranking system was devised based on contributing drainage area, with contributing area treated as a score that was discounted if the wetland area was >2.0% of the contributing area. Also, wetland sites with buffer areas more than twice the size of the wetland had their ranking score discounted. This preliminary ranking scheme illustrates that wetland sites could be prioritized to favor those sites that intercept larger contributing areas but remove relatively small areas of cropland from production. This scheme assumes that the nitrogen load reductions of the wetlands are captured by the criteria, an assumption that was tested through the following modeling exercise. 2.4. Modeling wetland loads and removal efficiencies The effectiveness of nitrate removal in a wetland can be estimated based on loads of nitrate and water entering the wetland (Crumpton et al., 2006; Dale et al., 2010). The USDA AnnAGNPS watershed simulation model (Bingner and Theurer, 2009) was used to estimate these loads for Lime Creek and at each identified wetland site. The AnnAGNPS model has been developed as a tool to evaluate pollutant loadings within a watershed and the impact of farming and conservation on hydrology and nonpoint pollution. Through continuous simulation of surface runoff, tile drainage, sediment, and non-point pollutant loads from watersheds, AnnAGNPS can be used to evaluate the impact of conservation practices on water quality. AnnAGNPS requires daily climate data to simulate watershed responses to rainfall and weather conditions. Spatial variability of soils, land use, and topography is accounted for by dividing the watershed into homogeneous cells from which the runoff and pollutants are routed downslope and then along simulated channels. Discharge can be simulated at field scales from precipitation inputs that can include rainfall, snowmelt, and irrigation. A daily soil water balance is maintained that can simulate tile drainage (Yuan et al., 2006), allowing discharge to include both surface and subsurface flow responses to precipitation. Sheet and rill erosion from each field is predicted based on the RUSLE model (Renard et al., 1997) and ephemeral gully erosion has recently been incorporated within AnnAGNPS (Bingner et al., 2010b). The model can be used to examine the effects of implementing various conservation alternatives within a watershed (Yuan et al., 2006, 2008; Bingner et al., 2010a). Input topographic (Fig. 2), landuse, and soils data were characterized for use with AnnAGNPS to simulate hydrology and water quality of the Lime Creek sub-watershed of Big Bureau Creek Watershed. The LiDAR-derived DEM was used as the topographic dataset. The 2009 land use was derived from the Illinois Cropland Data Layer (USDA-NASS, 2009). Soil information was obtained from

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the USDA Soil Survey Geographic (SSURGO) database as archived and distributed by the National Cartography and Geospatial Center (Soil Survey Staff, 2004). The cultivated corn/soybean crops represented 65% of the watershed, plus urban areas (one small village, low-density housing, roadways, and farmsteads), pasture, forest, and other crops. A thirty-year period of simulated weather was run to obtain results. The watershed discharge was calibrated to average the approximate long-term average discharge observed at a BBC gauging station near Princeton IL (USGS, 2010), since 1981, on a depth equivalent basis. The BBC drains a larger area with geology, soils and land use similar to Lime Creek. Total nitrogen loading and hydrologic discharge delivered to each candidate wetland location and to the Lime Creek watershed outlet was simulated and tallied on a monthly basis. Most of the total N load is assumed to be in the form of nitrate. Goolsby and Battaglin (2001) reported that 83% of the total N load in the Lower Illinois River was discharged in the form of nitrate. This proportion is likely to increase moving upstream into small watersheds where aquatic ecosystems have decreasing opportunity (time) to take up mineral N and cycle it into organic form. Nitrate removal in the wetlands was estimated using a simple approach based on a compilation of wetland studies across the Midwest (Crumpton et al., 2006, 2008; USEPA, 2008). Removal of nitrate in a wetland can be estimated from the hydraulic loading to the wetland as well as the nitrate concentration of the water entering the wetland. The hydraulic loading is the volume of water discharged from the wetland’s upgradient watershed divided by the area of the wetland itself. Crumpton et al. (2006, 2008) reported that in Midwest wetlands, percent nitrate removal (%NR) of the incoming load could be predicted as a function of annual hydraulic loading (AHL): %NR = 103 × AHL−0.33

(1)

where AHL is expressed in meters. The %NR rates become attenuated as hydraulic load is increased. This equation can be used to estimate mass nitrate removal (MNR) as the product of %NR and mass N load (i.e., the product of AHL and flow weighted average nitrate concentration, FWA) to obtain MNR = 10.3 × AHL0.67 × FWA

(2)

Estimates of nitrogen removal by the candidate potential wetlands were made using this approach. The method is based on annual data, with seasonality of discharge and nitrate-N concentrations typical in the Midwest (e.g., see Tomer et al., 2008) being germane. Results from the Lime Creek candidate wetlands are reported as annual averages and standard deviation of nitrogen removal within wetlands. The candidate sites were ranked based on estimated N removal rates. A final exercise was conducted to evaluate the feasibility of installing a wetland at the watershed outlet, considering how site selection criteria would need to be adjusted to do so, and how N removal efficiencies of smaller wetlands distributed in the watershed compare to one large wetland at the outlet. 3. Results and discussion 3.1. Nutrient removal wetlands Eleven sites were identified as potential locations for nutrient removal wetlands through the process of testing impoundments and subsequent field review (Fig. 4). The field review resulted in rejection of one candidate site identified using the LiDAR data, where the contributing area cropland was flat and very near the maximum buffer elevation. During the field review one wetland site was also added to the list of potential sites by moving the

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Fig. 4. Candidate wetland sites identified through terrain analysis and field review. Wetland identifiers are ranks given in Table 1, and assigned as described in the text.

simulated impoundment up-gradient about 150 m, to be above the ditch’s confluence with an ephemeral waterway and thereby avoid impeding drainage at a nearby roadway. During the field review every virtually every road-channel crossing was visited and no other rejected sites were deemed to be inappropriately excluded. This result emphasizes the success of the screening method, but also the importance of field checking candidate sites to verify their suitability for possible wetland installation. The eleven potential wetland sites were all located in the central and north-central part of the watershed (Fig. 4). This part of the watershed is the transition from the terminal moraine forming Lime Creek’s northern divide to the glacial-lacustrine plain in the southern part of the watershed. The lower part of the watershed had no sites that could be impounded under the selection criteria. The eleven candidate wetland sites were sorted into a preliminary ranking, according to size of contributing area (Table 1), after discounting where relative wetland and/or buffer areas were too large as previously described. In total, the potential wetlands could provide nutrient removal services for 30.3% (1976 ha) of Lime Creek watershed. The eleven wetlands sum to 38.9 ha, which, including buffer areas, would remove 122.3 ha from production. If only the lowest wetland site along each tributary is selected (this eliminates sites 5, 8, and 11, see Table 1), then the remaining wetlands would

occupy 25.1 ha, or 84 ha including wetland buffers. Installation of all the sites would result in 6.2% of the watershed to be taken out of production, while eliminating wetlands that occurred in series (i.e., omitting 5, 8, and 11) would lead to 4.2% of that portion of the watershed serviced being converted to wetlands. Note buffer areas were delineated conservatively; buffers could be managed for forage, biomass energy, or game bird habitat to continue to provide the landowner some form of economic return. 3.2. Wetland hydrologic and N loads Thirty years of simulated weather data plus soils and land-use information were used within the AnnAGNPS model to simulate hydrologic and nutrient loads discharged from the watershed. Average annual precipitation in the simulated weather record was 966 mm (38.0 in.), consistent with long term averages recorded for the area (Illinois State Water Survey, 2002, 2007). Average annual discharge simulated for the Lime Creek outlet was 279 mm, which was also consistent (i.e., within 10%) with measured data at the nearby BBC gauging station (gauge no. 05556500; USGS, 2010), which drains a larger area of the same glacial landform and averaged 300 mm annual discharge from 1981 to 2010. The total N

Table 1 Potential wetland sites, sorted according to contributing area and sizes of wetland and buffer areas (see text). Wetland locations are indicated in Fig. 2. Under the ‘Downgradient’ column, ‘yes’ denotes locations that are furthest down-gradient along a tributary channel (i.e., do not drain into another wetland). Contributing area (ha)

Wetland area (ha)

Wetland plus buffer area (ha)

Wetland site ID

Down-gradient

518 334 266 237 260 171 152 152 148 152 118

9.40 4.42 1.73 3.36 7.26 2.54 1.81 2.14 0.85 0.97 4.38

40.04 13.66 4.83 7.59 19.71 7.95 4.91 5.58 2.39 2.95 12.72

1 2 3 4 5 6 7 8 9 10 11

Yes Yes Yes Yes No Yes Yes No Yes Yes No

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Table 2 AnnAGNPS simulation results for annual hydraulic and nutrient loads at each of the potential wetland location and at the watershed outlet. Notes: Hydraulic load to the wetland (AHL) is equivalent to the contributing area discharge (Q) times the ratio of contributing area to wetland area. The same ratio converts N loss to N load. s.d.: standard deviation; FWA is arithmetic average of annual flow weighted average values. Wetland site ID

1 2 3 4 5 6 7 8 9 10 11 Outlet

Contributing area

Wetland

Q (s.d.) (mm yr−1 )

N loss (s.d.) (kg ha−1 )

AHL (s.d.) (M yr−1 )

293 (106) 279 (105) 267 (101) 302 (108) 272 (104) 270 (106) 303 (108) 231 (100) 299 (108) 261 (103) 314 (108) 279 (103)

31.5 (8.5) 24.7 (6.4) 28.9 (9.4) 44.4 (19.4) 21.2 (5.6) 13.3 (3.1) 30.9 (8.9) 5.5 (1.6) 35.5 (10.9) 17.9 (5.1) 39.1 (11.3) 28.6 (7.1)

16.4 (5.9) 21.7 (8.1) 44.9 (16.9) 21.0 (7.5) 10.9 (4.2) 19.2 (7.5) 25.3 (9.1) 17.4 (7.5) 55.1 (20.0) 42.4 (16.7) 8.5 (3.0)

load discharged on an average annual basis was simulated to be 28.6 kg ha−1 , which corresponded with a flow weighted average total N concentration of 10.2 mg L−1 . This result is also reasonable, and consistent with a BBC watershed assessment (Adamson and Dykstra, 2006). Multi-year studies of water quality in tile drained watersheds in Illinois (David et al., 1997; Kalita et al., 2006; Gentry et al., 2009) and Iowa (Schilling and Zhang, 2004; Schilling and Spooner, 2006; Tomer et al., 2008) have reported average NO3 -N loads >25 kg ha−1 yr−1 . Hydrologic discharge and total N loads delivered to each potential wetland site were also simulated (Table 2). These loadings varied among the locations, particularly for N loads, when expressed on the equalized basis of per unit area of contributing watershed. The variation is largely due to land use within the contributing area, as there are areas of pasture and low-density housing in the most northeastern part of the watershed, which drains to wetland sites 6, 8, and 10 (Fig. 4). The contributing area for site 8 has the least discharge and N load due to perennial vegetation associated with these land cover types (Table 2). Variation observed among the other sub-watersheds is due to effects of cropping pattern and land cover included in the model simulations, with some influence of soil type and slope on partitioning of runoff and sub-surface (tile) flows. When the variation among these wetland sub-catchments is combined with variation in the ratio of contributing to wetland area, the result is a greater than six-fold variation in wetland hydraulic load (Table 2). The range in variation in N loads delivered to the wetland sites is even greater than for wetland hydraulic load, with almost a 16-fold variation. These loading results are not correlated with the ranking obtained using size of contributing area as the main criteria. Wetlands would be best constructed where they have the opportunity to remove large N loads on a unit-area basis. While N removal efficiencies will inevitably decrease with larger hydraulic loads, mass N removal will generally increase with load as long as biogeochemical conditions are conducive to denitrification (i.e., availability of carbon and nitrate in an oxygen limited environment). Observations on nutrient removal wetlands in the Midwest (Crumpton et al., 2006; Dale et al., 2010) suggest that these conditions are well supported in wetlands receiving hydraulic loadings at least up to 60 m yr−1 . 3.3. Wetland N removal Estimates based on Eq. (1) (Crumpton et al., 2006; Dale et al., 2010) suggest that wetlands would remove 27–51% on the incoming nitrate load. However, the differences in estimated mass N removals provide the best comparison. The potential wetland sites with the largest estimated mass removal rates are those with the

N load (s.d.) (kg ha−1 ) 1757 (474) 1923 (501) 4843 (1586) 3084 (1345) 849 (222) 952 (225) 2588 (742) 410 (121) 6535 (2015) 2898 (823) 1062 (308)

FWA (s.d.) (mg L−1 ) 11.4 (2.9) 9.4 (2.6) 11.4 (3.2) 15.5 (5.7) 8.4 (2.4) 5.4 (1.6) 10.8 (2.8) 2.6 (0.9) 12.5 (3.6) 7.4 (2.4) 13.0 (3.3)

largest hydraulic loading. That is, wetland sites 3, 4, 7, and 9 are the four with the largest nitrate removal rates, while wetland sites 5, 6, 8, and 11 have the four lowest nitrate removal rates (kg ha−1 ). The preliminary ranking of wetland sites, largely based on contributing area, did not capture their relative N removal performance potential. Rather, N removal was dominated by relative N loading to the wetland, determined through hydraulic loading rate (related to contributing to wetland area ratio) and fraction of contributing area in row crops. The potential contribution of these wetlands toward watershed nutrient reduction was estimated for two implementation scenarios. The first one included wetland construction at the eight sites that were lowest (i.e., furthest down-gradient) along the tributary drainage ditches. This scenario omitted wetland sites 5, 8, and 11 (which drain to sites 1 and 2; see Fig. 4). The second scenario included wetland construction at all eleven sites identified through the LiDAR analysis. The second scenario required N removal rates for wetland sites 1, 2, and 5 (given in Table 3) to be adjusted because incoming N loadings were reduced by wetland sites located upgradient. Specifically, this involved reducing N loads to wetland 1 due to N removal in wetland 11, to wetland 5 due to N removal in wetland 8, and to wetland 2 due to removal in wetlands 8 and 5 (see Fig. 4). Under the first scenario, 25.1 ha of wetlands located at the eight furthest down-gradient wetland locations would reduce the watershed nitrate loss by an annual average of 11% (851 kg N ha−1 of wetland, with a standard deviation of 169 kg ha−1 yr−1 ). The second scenario, which increases the wetland area from 25.1 to 38.9 ha and includes all eleven sites, would reduce the watershed nitrate load by 13% (average of 637 kg ha−1 yr−1 of wetland with at standard deviation of 125 kg ha−1 yr−1 ). Note that wetlands located using the LiDAR data could provide up to a 13% reduction in watershed nitrate loss while only directly servicing 30% of the watershed. This is, proportionally, a 43% decrease from that area the wetlands would serve, which is approximately the 45% goal for the Mississippi River basin goal for hypoxia abatement. Wherever wetland sites can be located further down-gradient and remain adequately sized, a fewer number of wetland constructions might be required to achieve N reduction goals. However, implementing this strategy would require flexibility in siting criteria. We explored this issue at the Lime Creek outlet, and found it should be possible to install two wetland pools near the outlet that comprise 70 ha of pool area (Fig. 5), or 1.1% of the watershed area. If this could be done with only a 1 m elevation buffer, then a minimum of 144 ha of total land area (2.2% of the watershed area) might be needed to install this two-pool wetland. However, to construct the upper pool, local roads and one bridge

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Table 3 Estimated percent nitrate reduction for eleven simulated wetland sites. Mass nitrate reduction (MNR) estimates are given per unit area of wetland, based on Eq. (1) (Dale et al., 2010). Wetland sites are identified in Fig. 3. Wetland site ID

AHL (m)

N load (kg ha−1 yr−1 )

1 2 3 4 5 6 7 8 9 10 11

16.4 21.7 44.9 21.0 10.9 19.2 25.3 17.4 55.1 42.4 8.5

1757 1923 4843 3084 849 952 2588 410 6535 2898 1062

a

MNRa (kg ha−1 yr−1 ) 722 (130) 721 (128) 1423 (324) 1164 (409) 400 (74) 373 (59) 919 (180) 167 (43) 1795 (404) 873 (185) 540 (108)

%NR

Rank

41 37 29 37 47 39 35 40 27 30 51

6 7 2 3 9 10 4 11 1 5 8

Standard deviation of annual average values (based on 30-year simulation) given in parentheses.

would need to be reconstructed, and either raised or moved to an alternate route. The lower wetland of 21 ha, by itself, would not impact existing infrastructure but would have an average hydraulic loading of about 90 m/yr, requiring extrapolation beyond the range of data used to estimate and validate Eq. (1) (Crumpton et al., 2006; Dale et al., 2010). Based on Eq. (1) and AnnAGNPS-simulated loads, this 70 ha wetland should remove an average 960 kg NO3 N ha−1 yr−1 (s.d. = 157 kg NO3 -N ha−1 yr−1 ), from an average load of 2742 kg NO3 -N ha−1 yr−1 , which is a load reduction of 35% for the watershed. This is a substantial reduction from a single installation, but comes with considerations of increased planning procedures, public communications, and construction costs that would accompany a project of the larger scale. 3.4. Implications and alternative approaches Results of this study have several implications for watershed planning, conservation incentives, and nutrient trading schemes. Wetlands are clearly a key practice for reducing nitrate loads, and there is a need to understand the potential role that wetlands

can play in achieving N reduction goals in watersheds at a variety of scales. This exercise illustrates an approach that can be used to identify potential sites where wetlands can provide nitrate reduction in small, tile drained watersheds. Certainly, the placement criteria used in this study, particularly buffer height, should be viewed flexibly, and the setting and risk involved with any impoundment for wetland construction should be analyzed using a well-engineered approach. The buffer height of 1.5 m used here is regarded as very conservative. If the nitrate loads from Lime Creek had to be reduced by 45% to contribute to MRB nutrient reduction goals, then the eleven wetland sites identified across the watershed could provide about one fourth of that reduction. The flat terrain in much of the watershed limits the number of wetland sites that meet the applied criteria. It appears technically feasible to install a two-pool wetland near the outlet of the watershed that could, on its own, reduce the N load from the watershed by 35%. This large wetland would impact road infrastructure and face a more arduous administrative process. Therefore, alternative practices that could contribute to N-load reduction warrant consideration, including woodchip

Fig. 5. A two-pool wetland that could be installed near the outlet of Lime Cr. under less conservative criteria. Roadways crossing the upper wetland site would need to be rebuilt by being raised or re-routed.

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bioreactors (Schipper et al., 2010), controlled drainage (Drury et al., 2009), cover crops (Singer et al., 2011), and drainage ditch redesign (e.g., two-stage ditches; see Powell et al., 2007) deserve consideration. Part of the evaluation of alternative practices should include other ecosystem services provided, including carbon sequestration, biodiversity and wildlife habitat, and hydrologic function. Costs and feasibilities of a variety of practices might depend on the specific watershed, but each have been shown effective or have the potential to reduce nitrate loads in Midwest tile drainage. More research is needed to understand how different practices can be used in combination to meet specific water quality goals. Cover crops, controlled drainage and bioreactors may be appropriate in the flattest terrain where wetlands are not feasible, including the glacial-fluvial plain of the Lime Creek watershed. In larger watersheds with wider alluvial valleys and meandering streams, alluvial wetlands that receive diverted stream water off channel for nutrient removal would be an option to consider. This approach has been evaluated in European river basins (Koskiaho et al., 2003; Moreno-Mateos et al., 2010), and Mitsch and Day (2006) present a strategy for implementing this approach in the Mississippi River basin. Additional wetland sites in Lime Creek might be found by relaxing the criteria used in this study (i.e., in addition to the outlet), but this would need to be explored by evaluating alternatives using criteria that include social acceptability. If broader implementation of wetlands were found feasible, it should be possible to approach or exceed a 45% N-load reduction in Lime Creek with wetlands alone, given results of this study. Nutrient trading schemes are one approach for incentivizing all conservation practices that can reduce nutrient loads. This study indicates that individual wetland effectiveness in this watershed covered an order of magnitude (or a seven fold range in variation based on the updated method). Hypothetically, if one kg of N reduction is worth $9(USD), then the wetlands identified by LiDAR analysis would have nutrient-trade values between $1500 and $16,000 per ha of wetland per year. The potential 70 ha wetland at the outlet would have a nutrient trading value near the middle of this range, $8640 ha−1 . Therefore, purchasers of nutrient credits will need to be aware of this potential variation, which is based on expected N loads to wetland from land use in that specific watershed. Targeting wetlands at larger contributing areas should integrate and hence reduce this variation in performance.

4. Conclusions 1. LiDAR data were successfully used to identify suitable sites for wetland constructions in the Lime Creek watershed. Based on conservative criteria, approximately 30% of the watershed could be serviced by nutrient removal wetlands while converting 1–2% of cropland to wetland pools and converting another 3–4% to wetland buffer. 2. Modeling results suggest these wetlands would reduce annual N loss from the watershed by 11–13% based on two implementation scenarios (allowing or disallowing wetlands in series). 3. Technically, a two-pool wetland of 70 ha could be installed at the watershed outlet that would reduce the watershed N load by an estimated 35%. However, this installation would impact local road infrastructure. 4. Wetlands intercepting nutrients from relatively small contributing areas (100–500 ha) may be prone to widely varying nutrient removal efficiencies, depending on hydrology, land use, N loss from the contributing area, and size dimensions of the wetlands. This variation will have important consequences for achieving watershed N reduction goals and be a key consideration in

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marketing wetland sites for N removal credits under nutrient trading schemes.

Acknowledgements This research was funded in part by the US Environmental Protection Agency under a 2009 Targeted Watershed Grant. Thanks to Jared Bean for assisting with GIS processing of the LiDAR data.

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