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Effectiveness of Saturated Buffers on Water Pollutant Reduction from Agricultural Drainage
Gabriel Johnson1,*, Laura Christianson2, Reid Christianson3, Morgan Davis4, Carolina Díaz-García2, Tyler Groh5, Thomas Isenhart1, Jeppe Kjaersgaard3, Rob Malone6, Lindsay A. Pease7, Natalia Rogovska6
Published in Journal of Natural Resources and Agricultural Ecosystems 1(1): 49-62 (doi: 10.13031/jnrae.15516). 2023 American Society of Agricultural and Biological Engineers.
1Natural Resource Ecology and Management, Iowa State University, Ames, Iowa, USA.
2Crop Sciences, University of Illinois, Urbana, Illinois, USA.
3Minnesota Department of Agriculture, Saint Paul, Minnesota, USA.
4School of Natural Resources, University of Missouri, Columbia, Missouri, USA.
5Ecosystem Science and Management, Penn State University, University Park, Pennsylvania, USA.
6National Laboratory for Agriculture and the Environment, USDA ARS, Ames, Iowa, USA.
7Soil, Water, and Climate, University of Minnesota, Saint Paul, Minnesota, USA.
*Correspondence: gjohnson@iastate.edu
The authors have paid for open access for this article. This work is licensed under a Creative Commons Attribution-NonCommercial-NoDerivatives 4.0 International License https://creative commons.org/licenses/by-nc-nd/4.0/
Submitted for review on 30 December 2022 as manuscript number NRES 15516; approved for publication as a Review Article and as part of the Agricultural Conservation Practice Effectiveness Collection by Community Editor Dr. Ruth Book of the Natural Resources & Environmental Systems Community of ASABE on 3 June 2023.
Any opinions, findings, conclusions or recommendations expressed within do not necessarily reflect the view of the USDA. Mention of companies and manufacturers’ names does not imply endorsement by the USDA or coauthors. The USDA is an equal opportunity provider and employer.
Highlights
Abstract. It is a pivotal time in the development of saturated buffers as a conservation drainage practice. Field data have demonstrated that this practice can effectively reduce nitrate loads in subsurface drainage. The compilation and assessment of current knowledge for this relatively new practice is timely to help identify future opportunities. This review summarizes the state of the science for saturated buffers in the US within the context of this special collection’s emphasis on performance and cost. Suggested research areas are identified to improve understanding of saturated buffer function and performance and to refine design processes and criteria to accelerate adoption. As currently designed, saturated buffers removed an average of 46 ± 24% (mean ± sd) of the N load that would have otherwise entered receiving waters (9.4 ± 5.9 kg N removed/ha-y; n = 30 site-years). Cost efficiencies, which generally trended around $3 to $5/kg N removed per year, were considered relatively efficient compared to similar nitrate removal practices (range: $1.20 to $9.20/kg N/y), with planning level costs between $25 and $66/ha treated/y. As adoption is scaled, engineering design costs need to be considered unless the design model can be simplified. Future research should refine design processes, management, and siting criteria to facilitate scaled adoption for water quality goals. Additional studies on nutrient cycling within saturated buffers are needed to fill gaps about nitrogen and phosphorus dynamics and the role of buffer vegetation. Saturated buffers have significant nitrate reduction potential for tile-drained landscapes, but design adaptations may be needed to facilitate adoption in varied landscapes.
This article is part of a collection that provides a comprehensive review and evaluation of the performance and cost-effectiveness of selected agricultural conservation practices on nutrient and sediment reduction.
Subsurface drainage is a crucial agricultural water management practice for high-productivity agriculture in the Midwest. Yet, it comes with tradeoffs specifically with respect to water quality and quantity effects during critical seasons (Fausey et al., 1995; Dinnes et al., 2002). The saturated buffer is a conservation drainage practice designed to address these water resource concerns. Saturated buffers intercept and reroute subsurface (“tile”) drainage water to flow as shallow groundwater through a riparian buffer to reduce drainage nitrate loads. This practice reconnects the hydrology of tile-drained agricultural lands with adjacent riparian buffer soils by diverting discharge into a subsurface distribution pipe oriented parallel to the stream (Jaynes and Isenhart, 2014). A water control structure diverts the discharge, allows for bypass during heavy flows, and raises the water table in the saturated buffer into organic-rich soil horizons to promote denitrification and plant uptake (fig. 1). This practice has also been known as a saturated riparian buffer or a vegetated subsurface drain outlet, and it can be placed in a riparian buffer (forest or herbaceous) or in an edge-of-field filter strip.
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| Figure 1. Saturated buffer operation diagram. Tile flows are routed through the control structure, with the diverted flow saturating buffer soils. Interactions between nitrate (red diamonds) and microbes contribute to denitrification and conversion to nitrogen gas (purple circles). |
Traditional riparian buffers effectively reduce nitrate concentration in groundwater and improve water quality in receiving streams (Groh et al., 2020; Mayer et al., 2007; Spruill, 2000). Based on this concept, Jaynes and Isenhart (2014) designed a study to investigate rerouting a fraction of field tile drain flow as subsurface flow through a riparian buffer to increase nitrate removal. These early saturated buffer studies in central Iowa, USA, led to the development of a United States Department of Agriculture Natural Resources Conservation Service (USDA NRCS) Interim Conservation Practice Standard (CPS), which was adopted by several states beginning in 2012 (Interim CPS 739: Vegetated Subsurface Drain Outlet). Knowledge gained through implementation of this interim standard was used to create a new CPS in the NRCS National Handbook of Conservation Practices in 2016 (CPS 604: Saturated Buffer; USDA NRCS, 2016).
The CPS 604 provides requirements for saturated buffer design, construction, and management that are generally used in the USA. The NRCS defines a saturated buffer as a “subsurface, perforated distribution pipe used to distribute drainage system discharge beneath a vegetated buffer along its length and discharge channel.” The purpose is to “reduce nitrate loading from subsurface drain outlets through vegetation uptake and denitrification and to enhance or restore saturated soil conditions in riverine, lacustrine fringe, slope, or depression wetland hydrogeomorphic classes” (USDA NRCS, 2020). As per CPS 604, suitable sites for saturated buffers have the following characteristics: (1) soils with high organic matter content (>1.2% in the top 76 cm); (2) generally uniform hydraulic conductivity soils; and (3) a deep restrictive soil layer. These characteristics allow the water control structure to create the hydraulic gradient necessary to promote uniform lateral flow across the buffer to the stream (USDA NRCS, 2020). Optimal saturated buffer sites are designed to avoid possible negative impacts (e.g., reduced field drainage) of this managed outlet on adjacent crop fields. One of the practical advantages of this edge-of-field conservation practice is that water quality benefits can be achieved with minimal in-field disruption (Jaynes and Isenhart, 2014).
The overall objective of this paper is to summarize the state of the science for saturated buffers in the USA. Specifically, we synthesized the results of published studies to gain insight into practice effectiveness (pollutant loading reduction and cost) as well as identify research opportunities for design and implementation process improvement.
Review Method in Brief
A forward citation search starting with the original saturated buffer study of Jaynes and Isenhart (2014) was conducted to summarize the scientific literature on saturated buffers. An additional bibliometric analysis was performed via Web of Science (https://webofknowledge.com/) to identify citations using the Boolean strings ("saturat* buffer*" OR "riparian* buffer*") AND (tile* OR drain*) AND (nitr* OR no3*). Jaynes and Isenhart (2014) had 81 citations, followed by Tomer et al. (2015) with 55. The top two journal rankings were the Journal of Environmental Quality with eight articles and Environmental Monitoring and Assessment with three. These searches returned 58 articles from 2014-2022. Limiting the number of articles specifically to the conservation practice defined here and omitting more loosely related studies resulted in 26 articles, with only seven articles reporting performance effectiveness data from field studies. The remaining 19 articles primarily consisted of field studies focused on pollutant removal mechanisms and design considerations, and modeling studies focused on saturated buffer operation and projected water quality impacts from potential widespread implementation.
Performance Effectiveness
Nitrate-Nitrogen
Nitrate Removal Performance
Pooled across five studies including field data from five states (n = 47 site-years), saturated buffers removed an average of 63.7 ± 65.9 kg N/y (mean ± sd; median: 50.0 kg N/y; range: 0 to 300 kg N/y; table 1). This corresponded to a drainage treatment area-based removal of 6.4 ± 6.3 kg N/ha-y (median: 3.7 kg N/ha-y; range: 0 to 25.2 kg N/ha-y). The edge-of-field N loss reduction efficiency, calculated as the NO3 removed by the saturated buffer divided by the NO3 leaving the field in tile drainage water, averaged 37 ± 27% (median: 34%; range: 0% to 92%) (table 1). However, these 47 site-years included several sites that were later determined to be poorly designed or unsuitable due to landscape position and flow regime (Utt et al., 2015). These sites performed poorly for various reasons, including treating very little flow, monitoring interruptions due to flooding, insufficient carbon, and non-ideal placement; thus, removing those sites would give a more accurate estimate of reduction efficiency. Excluding data from these poorly performing early sites improved performance metrics to 46 ± 24% efficiency and 9.4 ± 5.9 kg N/ha-y (mean ± sd; n = 30 site-years; table 1, table A1). Notably, these 30 site years restrict the data to peer-reviewed studies.
| Table 1. Nitrate-N performance effectiveness in load reduction (%) at the edge-of-the field (including bypass flow), load reduction per contributing area (kg N/ha), and total N load removed (kg N) for saturated buffers. Full data (n=47) includes all monitored sites; Typical current designs (n=30) exclude saturated buffers deemed to be poorly performing in the early study by Utt et al. (2015). | |||||||
| Site Years[a] | Annual Impact on | Minimum | Reduction in Load | ||||
| 25th Percentile | Mean | Median | 75th Percentile | Maximum | |||
| Full Data (n = 47) | Percent Total N Load Reduction | 0 | 17 | 37 | 34 | 53 | 92 |
| N Load Removed per Contributing Area (kg N/ha) | 0 | 0.89 | 6.4 | 3.7 | 9.0 | 25.2 | |
| N Load Removed (kg N) | 0 | 12 | 64 | 50 | 82 | 300 | |
| Typical Current Designs (n = 30) | Percent Total N Load Reduction | 7 | 29 | 46 | 41 | 62 | 92 |
| N Load Removed per Contributing Area (kg N/ha) | 2.0 | 3.6 | 9.4 | 8.5 | 12.5 | 25.2 | |
| N Load Removed (kg N) | 13.0 | 46.3 | 83.4 | 61.5 | 114 | 300 | |
[a] Full site-year data in appendix (table A1). | |||||||
Two additional saturated buffer studies from Iowa were not included in these averages because flow was estimated using different techniques. All studies included for nitrate load reduction data in table 1 determined nitrate load from flow measured using common monitoring techniques of calibrated weirs in the saturated buffer control structures (see “Monitoring Recommendations” section), i.e., empirical measures of flow. In contrast, Streeter and Schilling (2021) estimated tile flow based on streamflow records and literature values of the tile flow fraction of streamflow and estimated the saturated buffer flow fraction of tile flow from data reported by Jaynes and Isenhart (2019a). Schilling and Streeter (2022) tested a paired riparian water table monitoring method to estimate hydraulic loading at the same saturated buffer. These two studies measured initial concentrations of 15 mg NO3-N/L being reduced to <1.5 mg NO3-N/L (90% concentration reduction in the buffer) for four site-years in eastern Iowa. Using the streamflow and literature method, Streeter and Schilling (2021) estimated load and loss reductions were 75 to 136 kg N and 12.3 to 13.5 kg N/ha, respectively, which were consistent with monitored values in this review (e.g., table 1) despite the different flow estimation methods. Using the paired water table monitoring method, Schilling and Streeter (2022) estimated load and loss reduction of 7 kg and 0.4 kg/ha, which were lower than most other published studies.
Nitrate removal performance within saturated buffers, be it defined as a percentage or on a load basis, depends on nitrate mass loading from the contributing area, the length of the distribution line, the hydraulic gradient, and the edaphic characteristics of the buffer. The interaction of these components is not fully understood, though it is apparent that most saturated buffers (as currently designed) effectively remove nitrate from the water that is routed into them. Here, the percentage of nitrate-N removed from diverted flow averaged 82 ± 22% across 29 site-years (median: 92%; range: 27%-100%; table A1). Much of this concentration reduction has been observed at wells adjacent to the saturated buffer distribution line, which suggests rapid N removal along the hydraulic gradient (Chandrasoma et al., 2022; Jaynes and Isenhart, 2019a; Streeter and Schilling, 2021).
With high N removal capacity within a saturated buffer, the total mass N load reduction of this edge-of-field practice is mostly limited by the fraction of drainage water diverted. A saturated buffer’s hydraulic capacity to receive flow can be exceeded during high flow periods in the spring/early summer, when the buffer is likely saturated (Chandrasoma et al., 2022). When hydraulic capacity is exceeded, the excess water is discharged directly to the stream. CPS 604 uses a flow capacity criterion where a saturated buffer’s design flow rate should be no less than “5% of the drainage system capacity or as much as practical based on the available length of the vegetated buffer” (USDA NRCS, 2020). Chandrasoma et al. (2022) illustrated uncertainties associated with estimating drainage system peak flow rates for use in edge-of-field conservation practice designs. Opportunities to optimize saturated buffer design to increase treated flow (i.e., reduce bypass flow) are discussed in the Future Research section.
Nitrate Removal Mechanisms
As with traditional riparian buffers, potential mechanisms of nitrate removal within saturated buffers are denitrification, plant uptake, and microbial immobilization (Groh et al., 2019a; Tufekcioglu et al., 2003; Groffman et al., 1992; Hill, 1996). The relative importance of each mechanism and potential interactions between mechanisms in a saturated buffer may depend on soil, climate, vegetation, and hydrology, and this is an area ripe for research as the practice of saturated buffers expands.
Saturated buffers are designed to facilitate denitrification by raising the water table of infiltrated water within the buffer into organic rich soil layers. However, elucidating denitrification as the specific nitrate removal mechanism remains a challenge due to the complicated biogeochemistry of the process. Groh et al. (2019a), using a modification of the static core–acetylene inhibition technique, reported that average denitrification rates for five saturated buffer site-years could account for between 3.7% and 77.3% of the total NO3 removed. The age of riparian buffers, with a likely corresponding increase in root density and labile soil organic C, positively correlated with denitrification rates (Groh et al., 2019a). This potential relationship was further demonstrated by Groh et al. (2019b), who showed that saturated buffer sites with limited in situ denitrification had a significant response to carbon additions in the laboratory setting. One potential mechanism for this relationship is that increased soil aggregation in older saturated buffers, especially microaggregates, can provide more anaerobic, C-rich microsites for denitrification and further increase denitrification rates (Seech and Beauchamp, 1988). As CPS 604 provides specific criteria on percent organic carbon and organic matter, further studies are warranted that elucidate nitrate removal mechanisms within saturated buffers and identify important controlling soil edaphic factors.
Although denitrification represents a permanent sink for NO3 as N2 gas in the atmosphere, incomplete denitrification could lead to the formation and release of N2O, a powerful greenhouse gas. The abundance of dissolved organic C in the saturated zone favors the production of N2, while N2O and NO production is more likely if microbial growth is restricted due to a small pool of soluble C (Kelso et al., 1999; Wang et al., 2013). In the only field study to date comparing total N2O emissions from saturated buffers, adjacent traditional riparian buffers, and cropped fields, Davis et al. (2019) concluded that the removal of NO3 by enhancing denitrification within saturated buffers did not result in an increase of N2O from incomplete denitrification. This study demonstrated that total N2O emission from saturated buffers, which included soil surface, dissolved, and indirect N2O emissions, over a period of two years, was no different from traditional riparian buffers at 3.85 kg N and 3.84 kg N, respectively. Total N2O emission from adjacent crop fields was ~3 times higher than that from the saturated buffer at 14.7 kg N. The results indicate that saturated buffers did not increase N2O emissions compared to traditional riparian buffers while substantially removing NO3 from reaching surface waters.
In general, N storage by buffer vegetation or immobilization by soil bacteria would temporarily remove the NO3 from the soil solution and store it as organic N. However, during decomposition, a fraction of this stored N would be mineralized and could re-enter the soil solution as inorganic N. Thus, uptake and sequestration do not represent a loss of N without a permanent removal of biomass and associated nutrients. Bosompemaa et al. (2021) demonstrated that during the growing season, nitrate concentrations in water in the vadose zone of a saturated buffer were similar between plots with riparian vegetation and bare soil plots. This was despite the vegetated plots having significantly lower soil nitrate concentrations. In this study, soil organic matter mineralization in bare plots contributed to higher soil NO3 content, while lower concentrations of soil NO3 in vegetated plots were attributed to plant uptake of mineralized N. This research highlights the complex nature of soil nitrogen dynamics, where nitrate uptake in the vadose zone is a short-term sink in which the plants continuously recycle the nitrogen, thus potentially affecting the timing of nutrient export.
Total Nitrogen
None of the saturated buffer studies evaluated reported total nitrogen (TN) information. Efforts to reduce nitrogen loading in Midwestern subsurface drainage focus on nitrate because this is the primary nitrogen form in subsurface discharge (Feth, 1966; Westcot, 1997; Randall and Goss, 2008). Specifically, Willrich (1969) reported that nitrate accounted for 99% of nitrogen in tile drainage discharge, and Baker et al. (1975) indicated that the ammonium fraction of tile drainage water at a field site in Iowa was less than 2%, with nitrate plus nitrite accounting for the remainder.
Sediment
A saturated buffer is not intended to function as a sediment removal practice for tile drainage flows. Subsurface drainage water tends to have lower suspended sediment concentrations than surface flows, but any sediment in the drainage water has the potential to settle in the control structure and/or restrict flow within the distribution pipes. The current practice standard for saturated buffers calls for avoiding surface intakes unless they are protected from debris entry (USDA NRCS, 2020). To date, Jacquemin et al. (2020) have provided the only saturated buffer study where sediment was assessed. They observed that inflow total suspended sediment concentrations in subsurface discharge (ranging from 0.5 to 28.4 mg/L) were completely removed within the buffer, as it was assumed infiltration through the buffer could not transport particulates.
While the primary objective of a saturated buffer is to reduce nitrate-nitrogen in subsurface drainage water, it is also expected that the buffer vegetation will function like a traditional riparian buffer or edge-of-field filter strip for surface flows. These traditional practices effectively slow surface runoff velocity, allowing sediment to settle within the buffer or filter strip (Lee et al., 2003; Dosskey, 2001). The extent of sediment removal in surface runoff water depends on factors including sediment particle size, buffer or strip width, type, density, stiffness of vegetation, and the presence of concentrated flow paths (Dosskey, 2001; Pankau et al., 2012).
In addition to slowing surface water runoff, riparian buffers increase soil infiltration rates when planted on ground that was recently a crop field. Bharati et al. (2002) found five times greater infiltration rates within riparian buffers when compared to adjacent farmland. However, the high water level (within 30 cm of the soil surface) in a saturated buffer could impact surface infiltration capacity. Studies on controlled drainage have shown an increase in surface runoff from the elevated drainage water level (Drury et al., 2009; King et al., 2022); however, this has not been assessed for a saturated buffer.
The sediment trapping and runoff reduction benefits of traditional riparian buffers and filter strips are well studied in many works, including Douglas-Mankin et al. (2021) and Yuan et al. (2009). Although it is expected that sediment reduction within saturated buffers would be similar to traditional buffers, this has yet to be explicitly documented. Additionally, potential impacts to surface infiltration due to conversion to perennial vegetation or due to the elevated water table from a saturated buffer are current gaps in the existing literature.
Streambank Stability
An initial concern about saturated buffers was the possibility for increased streambank sloughing due to the increased volume of water moving through the bank face. As a result, CPS 604 requires a slope stability analysis for sites with streambanks deeper than 2.4 m. Dickey et al. (2021) assessed this concern by estimating factors of safety for bank slope stability at five saturated buffers in Iowa and for 560 hypothetical saturated buffer scenarios representing a variety of designs and site conditions. From these simulations, the probability of bank failure increased by less than 3% due to the added subsurface flow due to the saturated buffer. Streambank height alone was not found to be a significant factor for streambank stability, with slope geometry (i.e., the combination of slope angle and bank height) having a larger effect. The analysis indicated that if a streambank was stable prior to saturated buffer installation, it would likely remain stable after installation, and thus it was concluded that bank height should not limit saturated buffer siting unless the potential site was already unstable. Additional empirical and modeling studies on the potential for bank sloughing would increase the robustness of these conclusions toward potential future updates of CPS 604.
Dissolved Phosphorus and Total Phosphorus
Phosphorus (P) is delivered to streams as particulate and dissolved P. Phosphorous transported through tile drainage is recognized as one of the main pathways of dissolved reactive phosphorus (DRP), contributing to total P loads in agricultural watersheds (Smith et al., 2015). For instance, DRP concentrations and loads in three tile-drained watersheds in Illinois were ~50% to 73% of total P concentrations (Gentry et al., 2007). However, few studies have analyzed phosphorus dynamics in saturated buffers, and the limited results indicate mixed performance. Utt et al. (2015) monitored total dissolved phosphorus at nine sites across Illinois, Iowa, and Minnesota over a two-year period and found no consistent trends in phosphorus removal. In Ohio, Jacquemin et al. (2020) monitored DRP at one site for one year and reported an 80% reduction of inflow concentrations ranging from less than 0.05 to 0.19 mg/L. However, Chandrasoma et al. (2022) reported no reduction in DRP concentrations over three years at three sites in Illinois, with the authors concluding that their bimonthly sampling resulted in too large of uncertainty to accurately assess annual phosphorus dynamics in the saturated buffers.
Particulate P is delivered to streams as a result of episodic surface runoff driven by excessive precipitation. Adoption of conservation practices, such as traditional riparian buffers designed to reduce or capture runoff, was shown to substantially reduce particulate P load to the watershed. Tsai et al. (2022) reported that the overall P removal efficacy of riparian buffers ranged from 51% to 55%. If saturated buffers are successful at reducing runoff sediment as suggested by Jacquemin et al. (2020), then it is likely that they will also reduce sediment-bound phosphorus loads. This could merit further investigation, particularly for watersheds that are affected by high sediment loads reaching tile lines and/or downstream sedimentation issues.
Cost Effectiveness
The cost effectiveness of saturated buffers was assessed only for nitrate ($/kg N removed). Limited saturated buffer phosphorus and sediment removal data precluded the development of cost efficiencies for those pollutants.
Jaynes and Isenhart (2014) developed the original cost efficiency of saturated buffers at $2.17/kg N/y based on an initial cost of $3500, a 20 y practice life, 4% interest rate, and annual N removal of 114 kg. Jaynes and Isenhart (2019a) later updated this to $2.94/kg N/y using a slightly higher initial cost of $4400, a longer practice life of 40 y, and a slightly more conservative removal of 73 kg N/y (range: $1.76 to $14.86/kg N/y for six sites). Although the NRCS has assigned saturated buffers a practice life of 15 years (USDA NRCS, 2019), there are no empirical bounds for practice life given the newness of this practice. Unlike other edge-of-field conservation practices (e.g., bioreactors), maintenance is relatively minimal as major maintenance is not required (e.g., replacement of carbon media). Additionally, in contrast to numerous in-field conservation practices that require specific annual management and expenses (e.g., cost of seeds for cover crop adoption), saturated buffers require no additional cost and only minimal annual maintenance. Maintenance operations consist of seasonal water level management and inspection of water control structures, and management of vegetation (e.g., mowing or removal of invasive species), which can all be done at the discretion of the landowner or operator.
The practice of saturated buffers was added as an approved practice in the Illinois Nutrient Loss Reduction Strategy in 2021, where an area-based cost efficiency of approximately $25/ha treated/y was reported. This was based on reported installation costs across literature, estimated maintenance (assumed $6.20/ha treated/y), an expected practice life of 20-50 y, drainage treatment areas of 10-21 ha, and a 6% discount rate (IEPA, IDOA, and University of Illinois Extension, 2021). Saturated buffer performance data from Illinois from Chandrasoma et al. (2022) has been newly used here with installation receipts to develop more precise cost efficiency estimates. In this new analysis, recurring annual costs were assumed for maintenance by the landowners (2 h/y at $40/h). Using a discount rate of 4% and a conservative practice life of 20 y yielded equivalent annual costs of $246/y, $368/y, and $457/y for three saturated buffers. This was higher than the $214/y reported in the assessment from Jaynes and Isenhart (2019a), possibly in part due to the inclusion of maintenance costs. Applying the monitored annual mass of N removed for the 10 site-years (34-300 kg N) resulted in cost efficiencies averaging $4.70 ± $2.80/kg N/y (mean ± sd; range: $1.20/kg N/y to $9.20/kg N/y). In terms of areal efficiency, these three saturated buffers were $25, $31, and $66/ha treated/y.
The development of multiple types of cost metrics (e.g., equal annualized costs in $/y, $/kg N/y, $/ha treated/y) allows practices to be compared as well as allows practice implementation to be prioritized and planned across larger scales. This comparison can be an important factor for estimating budgets for practice implementation at state or watershed-scales and for estimating levels of practice implementation necessary to achieve water quality goals. Cost estimates shown here for saturated buffers overlap with those found for other practices (fig. 2). For example, constructed wetlands have been reported to be $18 to $438/ha treated/y and $0.66 to $58/kg N/y (Messer et al., 2021), and denitrifying bioreactors have been reported to be $2.50/kg N/y to $20/kg N/y (Christianson et al., 2021a). The annual practice of winter cover cropping has cost efficiencies ranging from $72 to $162/ha treated/y and $1.50 to $5.00/kg N/y (Christianson et al., 2021b).
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| Figure 2. Cost range of winter cover cropping (Christianson et al., 2021b), denitrifying bioreactors (Christianson et al., 2021a), saturated buffers, and constructed wetlands (Messer et al., 2021) in (a) dollars per hectare treated per year (not reported for denitrifying bioreactors) and (b) dollars per unit mass of nitrogen removed per year. |
The Batch and Build approach in Iowa (discussed in more detail in the “Scaling adoption” section below) provided substantial evidence that engineering design costs are not immaterial when these practices are implemented at scale, but bundling individual sites together provided cost benefits from economies of scale. Estimates of total cost efficiency for the Batch and Build model to date have ranged from $2.76/kg N to $8.44/kg N when engineering expenses were included (assuming high and low treatment scenarios, 4% discount rate, 10 y life). If those services were “not taken into account,” cost efficiencies slightly improved to $2.05 to $6.24/kg N. Jaynes and Isenhart’s (2014, 2019a) early assumption that design costs would be minimal may be incorrect given the scale of saturated buffers called for across the landscape and given the current relatively intensive site survey and design methods. It is also worth considering if design processes could be simplified (e.g., to focus on distribution pipe length maximization and hydraulic gradient optimization) to reduce engineering costs, streamline saturated buffer implementation, and further improve cost efficiency.
A final component of cost efficiency is the potential for marketable biomass production from saturated buffer vegetation. Revenue from the harvest of buffer vegetation, whether for animal feed or bioenergy production, could help offset installation costs while aiding nutrient removal efforts. Future work on cost effectiveness and adoption should assess this component.
Monitoring Recommendations
Flow Monitoring
Monitoring saturated buffers for nitrate load reduction performance requires measurement of both flow volume and nitrate concentration. Common flow monitoring methods use three- or four-chambered water control structures with v-notch weir stoplogs and water level sensors in each upstream chamber to compute tile inflow, bypass flow, and flow to the buffer (fig. 3). The conversion of flow depth to flow rate requires the use of rating curves specific to the control structures used in the installation. These can be obtained from laboratory calibrations (Jaynes and Isenhart, 2014; Christianson et al., 2019; Shokrana and Ghane, 2021) or from water control structure manufacturers (Jacquemin et al., 2020). Pressure transducers have been the common choice for water level sensors (Jaynes and Isenhart, 2014, 2019a; Chandrasoma et al., 2022; Jacquemin et al., 2020); however, the authors have used other sensors, including radar and ultrasonic, which can offer benefits like increased accuracy and reduced drift but can often be more expensive or require specific working conditions.
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| Figure 3. Profile diagram of (a) 3-Chamber and (b) 4-Chamber water control structures showing flow paths and stoplog settings. Water level sensors are placed upstream of the v-notch weirs (orange lines). Qtotal = total flow in from field tile, Qbypass = bypass flow to receiving stream, Qbuffer = flow into saturated buffer. |
At least three chambers (two sets of stoplogs) are required in the control structure to fully monitor flow (fig. 3a). Inflow from the field tile is calculated by measuring discharge depth over the weir on the first set of stoplogs. Bypass flow is calculated from the discharge depth over the weir on the second set of stoplogs. The flow treated by the buffer is calculated as the difference between inflow and bypass flow (Jaynes and Isenhart, 2019a).
The accuracy of flow measurements can be enhanced by using a 4-chamber water control structure (three sets of stoplogs) or a splashguard plate to reduce turbulence at high flows (fig. 3b). In a 3-chamber structure, high flows can cause turbulence in the middle chamber, leading to flow “jumping” or bouncing over the second set of stoplogs and causing unreliable water level measurements. In a 4-chamber structure, the middle track of stoplogs is offset from the bottom of the control structure. This forces the flow from the first chamber to go down and back up before flowing over the third set of stoplogs, stilling the water before it flows over the bypass weir. A splash guard plate can be suspended in the middle chamber of a 3-chamber structure to retrofit existing installations to function like a 4-chamber structure.
It is recommended to verify sensor-derived water levels and discharges through regular site visits (e.g., weekly or biweekly during flow events) and manual measurements. Water level height may be measured using traditional tape measures, water-sensitive color changing paste, or electronic water level meters. Flow can be checked with a graduated bucket at the outlet and stopwatch measurements, if accessible. Additionally, stoplog adjustments should be recorded to ensure accuracy in weir calculations and aid in data validation. Furthermore, high flow observations should be documented. Reliable manual measurements may not be possible in high flow events, but documentation of event distribution and timing are vital for post processing. These observations should be recorded in a site notebook or a field log, such as in figure A1 of Christianson et al. (2021a), which could be modified for use with a saturated buffer.
An additional consideration when configuring monitoring of a saturated buffer is the potential for backwater or backflow conditions. This can happen when the water level in the stream rises and the water level in the downstream chamber of the control structure rises enough to interfere with free flow over the monitored weirs. In this situation, uncertainty surrounding flow into the saturated buffer as well as bypassing the saturated buffer are increased. There are limited solutions for this type of situation, though methods to monitor in-pipe water velocities may be an option. Existing studies have processed flow data to account for these conditions and provide conservative flow estimates (Chandrasoma et al., 2022).
Nutrient Concentrations and Loading
Nitrate concentration is usually monitored through weekly or biweekly (every two weeks) grab sample collection in the water control structure, monitoring wells in the buffer, and the adjacent stream. Sampling frequencies ranging from daily to biweekly are considered standard practice for monitoring nitrate in agricultural subsurface drainage water (e.g., Abendroth et al., 2022). Biweekly sampling provides an accuracy of ±10% for assessing the cumulative annual nitrate load removed because the nitrate concentration does not rapidly fluctuate in agricultural drainage water (Williams et al., 2015). To reach similar levels of accuracy for assessing cumulative DRP load, a much higher frequency of sampling is required. Williams et al. (2015) recommend water samples be collected every 13 to 26 hours for estimation of annual cumulative DRP load at an accuracy of ± 10%, while Dialameh and Ghane (2022) suggest using flow proportional sampling to increase accuracy and reduce the cost of cumulative P load compared to time proportional sampling.
Screened monitoring wells are generally installed in transects of two to four wells oriented from the saturated buffer distribution pipe towards the stream, with wells spaced approximately evenly across the buffer width (e.g., Chandrasoma et al., 2022, figures S1-S3; Jaynes and Isenhart, 2014, fig. 1; Jaynes and Isenhart, 2019a, fig. S2; Streeter and Schilling, 2021, fig. 1). The quantification of nitrate load reduction has been assessed by multiplying the difference in inlet and “outlet” concentration by the treated flow volume between sampling dates. Inlet concentration is defined as the concentration in tile inflow sampled in the water control structure. Because saturated buffers, as currently designed, do not have a single outlet (in contrast to other edge of field practices such as denitrifying bioreactors), the “outlet” concentration is assessed as the average across the wells nearest the stream. Jaynes and Isenhart (2019a) acknowledged that this method is unrefined and assumes uniform flow through the buffer; however, it is likely a conservative estimate. McEachran et al. (2023) provided further recommendations for monitoring well placement based on a three-dimensional, finite-difference groundwater model of saturated buffers. Based on the model, they recommended “centering the well screen on a line drawn between the distribution pipe and a point at about 75% of the mean water level in the stream,” (McEachran et al., 2023). This placement ensures the wells are capturing the primary path of tile flow through the buffer soils. While current monitoring well configurations have proven to be sufficient to monitor water table response to saturated buffer flow and nitrate concentration (Jaynes and Isenhart, 2014; Streeter and Schilling, 2021), caution should be taken in the design of research experiments to obtain accurate N loading by including representative proportions of concentrated flow paths, if present. Representing different soil textures proportionally across the buffer is one step that can be taken to maximize accuracy and minimize the number of wells installed.
Future monitoring efforts should prioritize (1) the use of 4-chambered control structures to reduce turbulence and the subsequent uncertainty of discharge rates; (2) the use of more intensive sampling frequencies to more accurately quantify P concentrations and load reductions; and (3) the possible development of a more robust and/or easier method to determine saturated buffer “outlet” nutrient concentrations.
Future Research: Opportunities to Scale Knowledge, Performance, and Adoption
Optimizing Design and Management for Improved Performance
Existing saturated buffer research has demonstrated that current designs effectively reduce N loss in subsurface drainage water. These studies have also made it clear that the overall mass load reduction is limited by the amount of flow treated. That is, increasing the amount of flow through (i.e., hydraulic loading of) saturated buffers has the potential to reduce the mass of N entering receiving waters even more than is currently achieved. The amount of treated flow is affected by soil saturated hydraulic conductivity, antecedent moisture conditions, the length of the distribution pipe, and the hydraulic gradient across the buffer. However, from a design perspective, the length of the distribution pipe and the hydraulic gradient across the buffer are the primary parameters that can be modified for a specific site.
Reducing bypass flow could be achieved by increasing capacity in the buffer via increased distribution pipe length. Indeed, the product of drainage area and distribution pipe length is one of the strongest predictors of saturated buffer nitrate-N mass removal (Jaynes and Isenhart, 2019a; Chandrasoma et al., 2022). However, increasing this length is not always feasible because the distribution pipe length can be limited by topography, property lines, or infrastructure.
Additional modifications to the distribution pipe beyond increasing length also offer potential options to improve saturated buffer flow capacity. Jaynes and Isenhart (2019b) modeled a dual-pipe configuration where distribution lines both at the field edge and midway between the field and the stream were simulated. This modification increased cumulative flow through the buffer by 3%-133% but was less effective for high permeability soils and under variable flow conditions (Jaynes and Isenhart, 2019b). Field studies are currently underway to assess dual line and pitchfork design styles (Illinois NREC, 2018; Kjaersgaard et al., 2016). Other potential modifications include the installation of multiple pipes at differing depths, the addition of multiple control structures to “step down” the distribution pipe, adding a gravel envelope around the distribution line, or sloping the distribution pipe away from the control structure. Based on a three-dimensional groundwater flow model of saturated buffers, McEachran et al. (2023) suggested designing the distribution pipe such that more flow leaves at the pipe ends away from the control structure to increase the travel time and thus increase nitrate removal.
The hydraulic gradient across the buffer is another controlling factor for the amount of flow treated, and this is a function of buffer width and the control structure stoplog settings (McEachran et al., 2020). A narrower buffer would support higher hydraulic gradients for a given hydraulic head and increase flow treatment capacity. While narrow buffers could risk reducing the residence time to be too low for nitrate removal, current studies (e.g., Jaynes and Isenhart, 2014; Jaynes and Isenhart, 2019a; Chandrasoma et al., 2022) have shown substantial nitrate removal typically occurs very close to the distribution line (e.g., within 5 m). In the only study to date assessing saturated buffer performance relative to width, McEachran et al. (2020) reported that optimal widths for six saturated buffers were all smaller than the constructed width, with two sites having smaller optimal widths than the Code 604 9.1 m minimum width. Thus, greater flexibility in the width criterion would enable more optimal designs. In the design process, the optimal saturated buffer width can be calculated using a theoretical methodology proposed by McEachran et al. (2020). This optimal width calculation, supported by improved estimates of the nitrate removal coefficient and saturated hydraulic conductivity, has considerable potential to improve saturated buffer designs for optimal flow treatment and overall effectiveness.
Other than modifying buffer width, stoplog height in the control structure could be raised to increase the hydraulic gradient across the buffer. However, this involves the management challenge of balancing the desire to increase saturated buffer flow treatment with the original agronomic objectives of subsurface drainage. That is, conservation drainage practices like saturated buffers should not restrict drainage capacity in the field, unless controlled drainage is specifically an aim. Fine-tuning this balance is a practical suggested research objective.
Stoplogs can be removed during planting and harvesting windows to allow full drainage capacity (e.g., Jaynes and Isenhart, 2014), but this results in all the discharge during those times bypassing treatment in the buffer. Recent developments in the real-time automation of control structure stoplog management could not only facilitate timelier and easier saturated buffer management but also improve conservation drainage practices in general. Automated control structures with internet-connected controls now allow producers to remotely program and change stoplog settings (e.g., Smart Drainage System, Agri Drain Corp., Adair, IA). This technological advancement should allow for more precise water management to meet both agronomic and conservation goals. Research on this is now ongoing (Burgis et al., 2022; Iowa Soybean Association, 2022).
International applications of saturated buffers offer a different design approach (Kjaergaard, 2022). A similar practice, an integrated buffer zone (IBZ), is being trialed in Europe, wherein both surface and subsurface runoff are collected in an edge-of-field pond (approximately 25 m long x 10 m wide) oriented parallel to a waterway. This pond is constructed deeper, closest to the field, to allow sediment to settle. The pond depth decreases nearer to the waterway, so the shallow side of the pond creates an inundated vegetated filter bed. Infiltrated water moves laterally from the pond towards the waterway, similar to groundwater flow in a saturated buffer. A low embankment can be constructed between the pond and the waterway to ensure sufficient residence time to provide nitrate removal through denitrification and plant uptake. IBZs have removed 10%-71% of total nitrogen from the water entering the pond (Zak et al., 2018; Carstensen et al., 2021). Zak et al. (2019) noted that in addition to water quality benefits, IBZs provide additional water storage, retention of sediment and particle-bound phosphorus from surface runoff, additional terrestrial and aquatic biodiversity, and biomass production from tree growth. Integrated buffer zones originated in Denmark and have since expanded to Britain, Sweden, Finland, and Germany (Zak et al., 2019).
Scaling Adoption Through Improved Siting
Widespread adoption of saturated buffers via state and federal programs has been relatively limited, despite edge-of-field practices being increasingly recognized as key to achieving water quality goals in tile drained areas (Feyereisen et al., 2022; The Nature Conservancy, 2021). A publicly available NRCS data dashboard indicates seven instances of Code 739 (Vegetated Subsurface Drain Outlet) and 19 instances of Code 604 (Saturated Buffer) being reported between 2013-2021 (USDA, 2021). These counts reflect only NRCS-assisted installations via federal funding, but nevertheless, this averages to a rate of less than three installations per year.
There is much room for saturated buffer adoption to increase, with Chandrasoma et al. (2019) estimating approximately 3.9 million ha across the US Midwest, or about 22% of the Midwest’s tile-drained area, with the potential to drain to a saturated buffer. Their simple GIS-based methodology considered characteristics such as buffer soil organic matter content >2.5% and the presence of poorly drained soils near the stream buffer area (i.e., fields likely to be tile-drained). Assuming 23% to 44% N loss reduction effectiveness of saturated buffers, N loading from tile-drained areas in 11 Midwestern states was projected to decrease by 5%-10% given full saturated buffer implementation.
Tomer et al. (2020) used the much more detailed Agricultural Conservation Planning Framework (ACPF) to identify specific areas for saturated buffers installation. In this study of 32 Iowa watersheds, 30% to 70% of streambank length was considered suitable for saturated buffer installation (i.e., not vulnerable to bank collapse or likely to cause inundation). Based on the estimated tile drained area upslope of these sites, tile drainage discharge from 15% to 40% of the watershed areas could be diverted. Unsuitable streambanks included the potential for bank failure due to saturation of high banks (>3.7 m) and/or inundation of adjacent crops upslope of the buffer due to slopes less than 2%. Also, headwater catchments covered a substantial area in some watersheds and were not suitable for saturated buffers. Notably, while Tomer et al. (2020) used a less conservative bank height restriction than CPS 604 (3.7 m vs 2.4), newer data from Dickey et al. (2021) demonstrated that the bank height criterion could be eased. Thus, these results from Tomer et al. (2020) are likely a conservative estimate.
These large-scale modeling efforts demonstrate the potential for saturated buffers to be adopted based on current knowledge of site acceptability criteria. Identification of such suitable sites could be streamlined and enhanced with novel geophysical techniques such as electrical resistivity and electromagnetic terrain conductivity imaging. These methods could be used to assess for the presence of undesirable characteristics such as sand or gravel lenses within a possible saturated buffer and provide more accurate and higher resolution data compared to traditional soil sampling and geotechnical investigation methods (Streeter and Schilling, 2021; Streeter et al., 2022).
Existing saturated buffer research has predominantly concentrated on sites in landform regions of the Wisconsinan glaciation (e.g., Schilling et al., 2022) due to the prevalence of poorly drained upland soils and the lateral flow conditions resulting from the underlying glacial till (Jaynes and Isenhart, 2019a). The adaptation of saturated buffers to function in other landform regions with different stratigraphy would provide additional opportunities for nitrate remediation. Modification of this practice for different landscape positions, such as in grassed waterways or in-field contour buffer strips, could also expand its beneficial use. These well-recognized existing conservation practices contain accumulated sediments conducive for denitrification, but their relative disconnection from N-rich drainage water presents an opportunity. Such “saturated waterways” would be a novel development on the current saturated buffer paradigm (Schilling et al., 2022).
Regardless of these possible siting criteria modifications to expand saturated buffer applicability in the future, current practice adoption needs to accelerate (e.g., Feyereisen et al., 2022). The Batch and Build approach in Iowa, a new implementation model for edge-of-field practices, has proven successful in accelerating the implementation and reducing total public expenditure for saturated buffers and denitrifying bioreactors (Kult, 2022). This new model aims to design and build dozens of saturated buffers and denitrifying bioreactors across a given watershed in one year. The two main features of this approach are: (1) the use of a fiscal agent to streamline financial transactions (e.g., the movement of funding from the state to engineers and drainage contractors) and (2) the combination of site selection, design, and construction of dozens of practices into a single “batch” to leverage efficiencies of scale. Kult (2022) reported that costs for saturated buffers in Iowa prior to the Batch and Build approach averaged $5800 for construction/materials plus $5700 for engineering design. With the Batch and Build model, engineering costs averaged $2900/practice for 51 practices (40 saturated buffers and 11 bioreactors).
Improving Performance Via Improved Mechanistic Understanding
Given the relative newness of the practice of saturated buffers, interesting biogeochemical questions remain (Groh et al., 2019a; Davis et al., 2019). How relative nutrient transformation pathways vary under different climates, flow conditions, and saturated buffer age is relatively unknown, especially given that the majority of current understanding is based on work in one state. Improved understanding of the nitrogen balance could also help quantify ancillary benefits, such as a reduction in nitrous oxide emissions compared to surrounding agricultural fields (Davis et al., 2019). These ancillary benefits likely extend beyond the nitrogen cycle to include wildlife and pollinator habitat and biomass production, but these more wholistic benefits have yet to be studied specifically at saturated buffers.
Considering the importance of denitrification in saturated buffer functioning, the thickness of organic matter-rich soil layers and soil carbon content are critical (Jaynes and Isenhart, 2014; Groh et al., 2019b; Hill and Cardaci, 2004). Burford and Bremner (1975) showed that soil with organic carbon contents of at least 2% can easily sustain high denitrification rates. The CPS 604 calls for saturated buffers to be placed on soils with at least 1.2% organic matter content (0.75 percent organic carbon) in the top 76 cm. Current research has established that saturated buffers designed under these criteria work, and future research could help assess performance sensitivity to these soil criteria with the possible aim of refining those values further. This could potentially expand suitable sites to those with lower organic matter contents or different stratigraphy, such as buried soil profiles from the placement of drainage ditch dredging spoil or from the deposition of eroded topsoil (Streeter and Schilling, 2021).
Buffer vegetation within a saturated buffer has also been shown to impact water quality improvement and provide possible ancillary benefits (Bosompemaa et al., 2021). Vegetation types possible for use in a saturated buffer (e.g., trees, woody shrubs, perennial grasses, pollinator mixtures, forest and herbaceous mixtures, and berry and nut crops) differ in their ability to assimilate nitrogen and phosphorus. They also differ in their marketable biomass productivity, which could be of interest to some landowners. Any biomass or crop harvested represents a permanent removal of nutrients from the buffer area, and these beneficial nutrient balances have not been assessed.
The positive correlation of buffer age with denitrification and nitrate removal implies a potential C limitation (Jaynes and Isenhart, 2019a; Groh et al., 2019a,b). However, optimization of the number and distribution of vegetative species within a buffer to maximize carbon availability and enhance nitrate loss is yet to be explored. Given in-field weed management and chemical use at many farms, there are also questions about the possible invasiveness of saturated buffer species as well as the vegetation’s herbicide resistance. One potential tradeoff is that raising the water table within the riparian zone may provide a competitive advantage to Reed canary grass (Phalaris arundinacea), which can form dense, dominant cover and require control when there are management goals related to maintaining or restoring other plant species. Finally, it is possible that deep rooted vegetation may enter and clog the distribution lines. While existing studies have not noted distribution line clogging (Jaynes and Isenhart, 2019a), research on root intrusion is ongoing (Iowa Learning Farms, 2022). Buffer vegetation is an essential saturated buffer design component, and the lack of studies in this area leaves much room to grow.
This review has demonstrated that there is a need to quantify P dynamics within saturated buffers. Many locations where a saturated buffer is a suitable practice have dual N and P loss reduction goals (e.g., Mississippi River Basin, Chesapeake Bay watershed). It is possible that improved or refined management of saturated buffers could aid in P removal. Reducing drainage discharge through the combined practices of controlled drainage paired with a saturated buffer could reduce P loss from a field. Selecting buffer vegetation to have high P uptake capacity and then harvesting the vegetation could be another management approach to address dual nutrient concerns. However, P balances are entirely lacking for saturated buffers to date.
Conclusions
Saturated buffers are an effective and cost-efficient practice for reducing nitrate-nitrogen loads from subsurface drainage discharge. Edge-of-field nitrate-nitrogen loss reductions average 46 ± 24% (mean ± sd) for sites designed for current practice standards. The cost efficiency of saturated buffers is similar to other edge-of-field conservation practices, ranging from $1.20 to $9.20/kg N/y. Monitoring protocols should include both flow and nitrate concentration measurements to quantify nitrate mass loads. Existing studies have documented significant nitrate loss reduction within saturated buffers and indicated that denitrification is the primary removal mechanism, but work remains to fully elucidate nitrogen biogeochemical processes and potential impacts on additional pollutants (e.g., phosphorus). Practice effectiveness is limited by hydraulic loading capacity, while implementation efforts can be enhanced through improved site suitability guidance (e.g., width, soil organic carbon, and bank height criteria).
Meeting water quality goals across the U.S. Midwest will require a multitude of land use, in-field, and edge-of-field practices. The saturated buffer is proven to be an effective edge-of-field nitrate-nitrogen loss reduction practice and should continue to be a key component of nutrient reduction strategies. Compared to other practices, saturated buffers provide advantages in terms of lower maintenance requirements (compared to in-field activities and denitrifying bioreactors) and lower costs (compared to wetlands). However, suitable sites are currently limited to riparian corridors meeting design criteria (e.g., carbon availability, soil type, topography). As implementation scales up, it is imperative for saturated buffer science to continue advancing. Thus, future research is recommended to focus on (1) design optimization and flow management to maximize treatment efficiency and nitrate load reduction; (2) expansion of suitable sites by investigating current limitations (e.g., carbon availability) as well as the suitability of additional landform regions and landscape positions; and (3) improved understanding of the relationships between soil edaphic properties and nitrate removal capacity. Optimization of hydraulic loading capacity can improve the effectiveness of current design practices to maximize nitrogen mass removal and cost effectiveness. A greater understanding of soil edaphic properties, carbon availability, and nitrate removal capacity could increase the number of suitable sites to those with different carbon amounts or stratigraphy. Adaptation of the practice to alternate landscape positions (e.g., the “saturated grassed waterway concept”) would also expand the potential for suitable sites. Additional investigation of effects on other pollutants, such as phosphorus, would provide information on potential co-benefits beyond nitrogen removal. Addressing these knowledge gaps will enhance the pollutant reduction performance and improve the cost effectiveness of saturated buffers, aiding efforts to protect and improve agricultural water quality.
Acknowledgments
The authors gratefully acknowledge funding from: USDA NRCS CIG (NR213A750013G038); Illinois Nutrient Research and Education Council (#2017-4-360498-168 and #2021-3-360498-144); and the Iowa Nutrient Research Center. The authors thank Chelsea Ferrie for creating the initial summary of the published data.
Appendix
| Table A1. Nitrate removal data of all site years for studies compiled in this review. Abbreviations and symbols are as follows: EOF = Edge-of-Field, * = water year (October – September), ^ = partial flow data for that year, and [a] = assumed value. Jaynes and Isenhart (2014) initially reported data from BC-1 for 2010 and 2011 but Jaynes and Isenhart (2019a) republished it, thus that data is attributed to the 2018 study here. | |||||||||||
| Study | State | Site | Year or Monitoring Period | Nitrate Load from Contributing Area (kg N) | Nitrate Yield from Contributing Area (kg N/ha) | Nitrate Load Diverted to Buffer (kg N) | Nitrate Diverted Per Contributing Area (kg N/ha) | Nitrate Load Removed (kg N) | Nitrate Removed Per Contributing Area (kg N/ha) | Buffer Removal Percentage | EOF N Loss Reduction Percentage |
| Jaynes and Isenhart, 2019a | Iowa | BC-1 | 2011 | 220 | 21.8 | 123 | 12.2 | 123 | 12.2 | 100.0 | 56 |
| Iowa | BC-1 | 2012 | 200 | 19.8 | 92.0 | 9.1 | 84 | 8.3 | 91 | 42 | |
| Iowa | BC-1 | 2013 | 514 | 50.9 | 181 | 17.9 | 179 | 17.7 | 99 | 35 | |
| Iowa | BC-1 | 2014 | 320 | 54.2 | 111 | 18.8 | 111 | 18.8 | 100 | 35 | |
| Iowa | BC-1 | 2015 | 233 | 39.5 | 85.0 | 14.4 | 80 | 13.6 | 94 | 34 | |
| Iowa | BC-1 | 2016 | 111 | 18.8 | 55.0 | 9.3 | 53 | 9.0 | 96 | 48 | |
| Iowa | BC-1 | 2017 | 151 | 25.6 | 51.0 | 8.6 | 47 | 8.0 | 92 | 31 | |
| Iowa | IA-1 | 2014 | 67 | 14.3 | 56.0 | 11.9 | 55 | 11.7 | 98 | 82 | |
| Iowa | IA-1 | 2015 | 76 | 16.2 | 70.0 | 14.9 | 59 | 12.6 | 84 | 78 | |
| Iowa | IA-1 | 2016 | 62 | 13.2 | 61.0 | 13.0 | 57 | 12.1 | 93 | 92 | |
| Iowa | IA-1 | 2017 | 41 | 8.7 | 40.0 | 8.5 | 36 | 7.7 | 90 | 88 | |
| Iowa | B-T | 2016 | 60 | 8.5 | 28.0 | 3.9 | 24 | 3.4 | 86 | 40 | |
| Iowa | B-T | 2017 | 37 | 5.2 | 23.0 | 3.2 | 23 | 3.2 | 100 | 62 | |
| Iowa | HG | 2016 | 222 | 10.2 | 47.0 | 2.2 | 46 | 2.1 | 98 | 21 | |
| Iowa | HG | 2017 | 225 | 10.3 | 65.0 | 3.0 | 64 | 2.9 | 99 | 28 | |
| Iowa | BC-2 | 2016 | 1758 | 43.4 | 444 | 11.0 | 120 | 3.0 | 27 | 7 | |
| Iowa | BC-2 | 2017 | 1334 | 32.9 | 234 | 5.8 | 115 | 2.8 | 49 | 9 | |
| Iowa | SH | 2016 | 61 | 17.4 | 27.0 | 7.7 | 13 | 3.7 | 48 | 21 | |
| Iowa | SH | 2017 | 205 | 58.6 | 94.0 | 26.9 | 31 | 8.9 | 33 | 15 | |
| Chandrasoma et al., 2022 | Illinois | SB1 | 2019* | 178 | 18 | 37.6 | 3.8 | 34 | 3.5 | 90 | 19 |
| Illinois | SB1 | 2020* | 112 | 11.3 | 84.2 | 8.5 | 82 | 8.3 | 97 | 73 | |
| Illinois | SB1 | 2021* | 148 | 14.9 | 81.2 | 8.2 | 68 | 6.9 | 84 | 46 | |
| Illinois | SB2 | 2018* | 95.9 | 13.9 | 76.6 | 11.1 | 59 | 8.6 | 77 | 61 | |
| Illinois | SB2 | 2019* | 179 | 26 | 79.4 | 11.5 | 50 | 7.3 | 63 | 28 | |
| Illinois | SB2 | 2020* | 343 | 49.7 | 205 | 29.7 | 117 | 17 | 57 | 34 | |
| Illinois | SB2 | 2021* | 185 | 26.8 | 121 | 17.5 | 66 | 9.5 | 55 | 36 | |
| Illinois | SB3 | 2019* | 388 | 32.6 | 309 | 26.0 | 300 | 25.2 | 97[a] | 77 | |
| Illinois | SB3 | 2020* | 431 | 36.2 | 239 | 20.1 | 232 | 19.5 | 97[a] | 54 | |
| Illinois | SB3 | 2021* | 331 | 27.8 | 164 | 13.8 | 159 | 13.3 | 97[a] | 48 | |
| Utt et al., 2015 | Iowa | IA-3 | 2015 | 639 | 10.7 | N/A | N/A | 185 | 3.08 | N/A | 29.0 |
| Illinois | IL-2 | 2014 | 886 | 34.9 | N/A | N/A | 133 | 5.24 | N/A | 15.0 | |
| Illinois | IL-3 | 2014 | 7.1 | 0.46 | N/A | N/A | 1.4 | 0.09 | N/A | 19.7 | |
| Illinois | IL-3 | 2015 | 108 | 6.94 | N/A | N/A | 30.8 | 1.99 | N/A | 28.7 | |
| Illinois | IL-4 | 2014^ | 45.8 | 6.59 | N/A | N/A | 38.1 | 5.48 | N/A | 83.2 | |
| Illinois | IL-4 | 2015 | 72.6 | 10.5 | N/A | N/A | 2.9 | 0.42 | N/A | 4.0 | |
| Illinois | IL-5 | 2014^ | 21.8 | 0.36 | N/A | N/A | 5.9 | 0.10 | N/A | 27.1 | |
| Illinois | IL-5 | 2015 | 689 | 11.4 | N/A | N/A | 73 | 1.21 | N/A | 10.6 | |
| Indiana | IN-1 | 2014^ | 44.4 | 14.9 | N/A | N/A | 0 | 0.00 | N/A | 0.0 | |
| Indiana | IN-1 | 2015 | 30.8 | 10.3 | N/A | N/A | 1.59 | 0.53 | N/A | 5.2 | |
| Indiana | IN-2 | 2014^ | 0.82 | 0.14 | N/A | N/A | 0.68 | 0.12 | N/A | 82.9 | |
| Indiana | IN-2 | 2015 | 35.4 | 6.25 | N/A | N/A | 1 | 0.18 | N/A | 2.8 | |
| Minnesota | MN-2 | 2015 | 73.5 | 3.60 | N/A | N/A | 11.8 | 0.58 | N/A | 16.1 | |
| Minnesota | MN-3 | 2014^ | 6.8 | 0.59 | N/A | N/A | 2.3 | 0.20 | N/A | 33.8 | |
| Minnesota | MN-3 | 2015 | 8.2 | 0.72 | N/A | N/A | 0 | 0.00 | N/A | 0.0 | |
| Minnesota | MN-4 | 2014^ | 29 | 1.79 | N/A | N/A | 5 | 0.31 | N/A | 17.2 | |
| Minnesota | MN-4 | 2015^ | 68 | 4.21 | N/A | N/A | 1.54 | 0.10 | N/A | 2.3 | |
| Jacquemin et al., 2020 | Ohio | N/A | June 2018 – June 2019 | N/A | N/A | N/A | N/A | 15 | 2.05 | N/A | 75.0 |
| Streeter and Schilling, 2021 | IA | N/A | 2019 | N/A | N/A | N/A | N/A | 82 to 136 | 12.3 to 13.5 | 90.0 | N/A |
| IA | N/A | June 2020 – December 2020 | N/A | N/A | N/A | N/A | 75 to 125 | 12.3 to 13.5 | 90 | N/A | |
| Schilling and Streeter, 2022 | IA | N/A | 2021 | N/A | N/A | N/A | N/A | 7 | N/A | N/A | N/A |
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