Groundwater nitrate reductions within upstream and downstream sections of a riparian buffer

Ecological Engineering 47 (2012) 297–307

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Groundwater nitrate reductions within upstream and downstream sections of a riparian buffer Tiffany L. Messer a , Michael R. Burchell II a,∗ , Garry L. Grabow a , Deanna L. Osmond b a b

Biological and Agricultural Engineering, Box 7625, North Carolina State University Campus, Raleigh, NC 27695, United States Soil Science, Williams Hall 3403D, Box 7619, North Carolina State University Campus, Raleigh, NC 27695, United States

a r t i c l e

i n f o

Article history: Received 20 December 2011 Received in revised form 23 May 2012 Accepted 22 June 2012 Available online 10 August 2012 Keywords: Riparian buffer Groundwater Hydrology Nitrate NO3 − -N

a b s t r a c t The objective of this study was to evaluate the water quality benefits provided by a buffer enrolled in the North Carolina Conservation Reserve Enhancement Program (NC CREP). A 5-year study was conducted on two distinct buffer sections along the same stream to evaluate the hydrology and attenuation of groundwater nitrate (NO3 − -N) entering from nearby agricultural fields. The average buffer widths were 60 m (Section 1, upstream) and 45 m (Section 2, downstream). Three transects of groundwater monitoring well nests within each buffer zone were installed to monitor water quality and water table depths for 5 years. Mean groundwater NO3 − -N concentrations at the 1.5 m depth decreased from 4.5 mg L−1 to 1.7 mg L−1 and from 12.9 mg L−1 to 1.4 mg L−1 in buffer Sections 1 and 2 respectively. These differences were significant in both buffer sections (˛ = 0.05), but the wider Section 1 received significantly less NO3 − -N than did Section 2 (P < 0.0001). Groundwater NO3 − -N loads were reduced by 0.003 kg m−2 yr−1 (76% reduction) at the 1.5 m depth, while in Section 2 these loads were reduced by 0.02 kg m−2 yr−1 (94% reduction) and 0.04 kg m−2 yr−1 (86% reduction) at the 1.5 m and 3 m depths, respectively. Topography, water table and redox measurements, nitrate to chloride ratios, and deep groundwater cation analyses, indicated both sections were suitable for denitrification to proceed. However, the position of the wider Section 1 buffer in the landscape limited the amount of NO3 − -N contaminated groundwater that entered from the agricultural fields, and thus could have been designed to be narrower. The effectiveness of NO3 − -N reduction in riparian buffer systems is dependent on multiple landscape and biogeochemical factors and not buffer width alone. Findings provide design guidance for conservation buffer program managers as related to the influence of buffer landscape position on groundwater nitrate reduction. © 2012 Elsevier B.V. All rights reserved.

1. Introduction Riparian buffers are important best management practices (BMPs) because they can remove pollutants from surface runoff and shallow groundwater prior to stream discharge. Defined as a complex assemblage of soil, plants, and organisms immediately adjacent to a water course that may include wetlands, stream banks, and floodplains (Lowrance et al., 1985; Osmond et al., 2002), riparian buffers have been found to reduce nitrate (NO3 − N) concentrations up to 90% from groundwater originating from

Abbreviations: BMP, best management practice; NO3 − -N, nitrate-nitrogen; NC CREP, North Carolina Conservation Reserve Enhancement Program; Cl, chloride; DOC, dissolved organic carbon; Na+ , sodium; Ca2+ , calcium; USACE, United States Army Corps of Engineers; FE, field edge; SE, stream edge; MB, mid buffer; redox, oxidation/reduction. ∗ Corresponding author. Tel.: +1 919 513 7372. E-mail address: mike [email protected] (M.R. Burchell II). 0925-8574/$ – see front matter © 2012 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.ecoleng.2012.06.017

an upland land use, with losses primarily attributed to microbial denitrification (Peterjohn and Correll, 1984; Lowrance, 1992). If the groundwater direction is predominately toward the buffer for the majority of the year, maximum NO3 − -N removal through denitrification can occur if the groundwater encounters suitable soil carbon levels, anoxic (i.e. low redox values) conditions, and moderate temperature and pH conditions (Korom, 1992; Puckett, 2004). Many studies have focused on the physical, hydrologic, and biogeochemical conditions that maximize NO3 − -N removal through denitrification in these systems (Gilliam, 1994; Schultz et al., 1995; Hill, 1996; Spruill, 2004; Evans et al., 2007). However, results are often variable with NO3 − -N reductions ranging anywhere from 0% to 95% (Spruill, 2004). The fundamentals of riparian buffer performance are understood well enough for conservation programs to encourage their use as BMPs in North Carolina to combat the effects of nonpoint source pollution from sources such as agriculture (NC DENR, 2004). Programs such as the North Carolina Conservation Reserve Enhancement Program (NC CREP) have developed an increasing

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interest in buffer placement. To encourage enrollments, the NC CREP provide a one-time bonus and yearly land rental payments to landowners, many of whom are farmers, to restore riparian and wetland areas to improve water quality (NC CREP, 2008). By the nature of these programs, riparian buffers are often designed for and implemented at sites with variable hydrologic and biogeochemical regimes. Removal of NO3 − -N is often variable, with many buffers failing to provide the level of treatment desired (Ocampo et al., 2006). To maximize the impact to water quality improvement, these programs have a vested interest in defining ideal buffer sites whose contribution to water quality improvement justifies the cost of land payments. Width and landscape position are the two major buffer characteristics that these programs consider when enrolling lands to be used as riparian buffers. Unfortunately, landowner preferences and pressure for these programs to meet acreage goals often lead to excessively wide buffers. These wider buffers lead to an extra cost to the program and reduction of agricultural production with little to no extra benefit to water quality improvement. The effects of increased buffer widths on NO3 − -N reduction have been studied to investigate the benefits of taking these areas out of agricultural production. A study by Dukes et al. (2002) on four relatively narrow riparian buffers reported that wider plots (15 m) decreased NO3 − -N levels 15% more than narrower plots (8 m), with differences attributed to increased residence times through the buffer. Mayer et al. (2007) estimated buffer NO3 − -N reduction through a meta-analysis of 89 buffers with variable widths. NO3 − -N reduction was found to significantly increase as widths increased from 0 to 25 m. However, increasing width from 25 to 50 m did not significantly increase NO3 − -N removal. Findings were attributed to higher water tables and carbon availability in buffer portions that were closest to the stream, resulting in more suitable conditions for denitrification to occur. Angier and McCarty (2008) studied wider riparian buffers that varied from 60 to 250 m. Topographic differences along the buffer’s field edge allowed one portion of the buffer to receive higher concentrations of NO3 − -N than upstream portions making comparisons difficult. The study concluded that NO3 − -N reduction is not only dependent on buffer width, but also groundwater flow direction and depth. Additional studies have also reported that ideal buffer placement is highly dependent on not only the physical dimensions of the buffer, but also on topographic location relative to adjacent pollutant sources, soil horizons, water table elevation, and dissolved organic carbon availability (Lowrance, 1992; Lowrance et al., 1995; Hill et al., 2000; Dukes et al., 2002; Vidon and Hill, 2004; Puckett and Hughes, 2005). Vidon and Hill (2004) modeled groundwater in eight riparian sites to identify topographic effects on NO3 − -N reduction and water table fluctuations. Topographic qualities in the riparian zones were critical components for decreasing runoff and groundwater velocity within the buffer. Decreased velocities from flatter topographies increased groundwater residence time within riparian buffers allowing more opportunities for denitrification. Hill et al. (2004) evaluated denitrification potentials with soil core samples from five riparian buffers to 400 cm depth. The study found denitrification activity in layers down to 210 cm. Furthermore, Hill et al. (2004) reported that NO3 − -N concentrations were lower due to increased denitrification rates in coarse sediment layers that were receiving carbon leaching downward from overlying organic rich horizons. Puckett and Hughes (2005) investigated the transport and fate of NO3 − -N in a shallow aquifer and adjacent stream of a dairy farm and reported hydrogeologic factors (i.e. tile drains, residence time, and coarse sediments) strongly influenced the transport and fate of NO3 − -N. Studies have shown how denitrification can occur in these systems and the mechanisms that can affect their pollutant reduction

efficiencies. However, research is still needed, particularly on sites that have been enrolled in conservation programs, to help managers understand the relationships between buffer width, location in the landscape with respect to source, site hydrology, and soil biogeochemical conditions that will maximize NO3 − -N removal. To maximize the impact riparian buffers can have on water quality improvement, these relationships should determine the locations and designs of new riparian buffers, rather than allowing widths to be determined arbitrarily by field shape or landowner preferences. In an effort to improve the future impact of the NC CREP, a detailed, 5-year evaluation of hydrology and groundwater NO3 − -N attenuation was conducted on two sections of a buffer enrolled in the program. These sections were located along the same first-order stream and received elevated levels of groundwater NO3 − -N from the same row-crop fields, but had two distinct widths and landscape positions. Research objectives were to (1) evaluate NO3 − -N removal performance of the two buffer sections, (2) determine the characteristics of the site (i.e. hydrology, width, landscape position) that may have impacted observed performance, (3) identify contributions of denitrification and dilution to observed NO3 − -N removal, and (4) provide recommendations about the guidelines that should be used in future NC CREP riparian buffer enrollments with similar soil series in the upper Coastal Plain, to maximize water quality improvements. 2. Materials and methods 2.1. Site description The study site was located on a row crop farm in Halifax County, north of Enfield, NC. The farm was situated in the upper Coastal Plain region of North Carolina within the Tar Pamlico River basin. The site was chosen because it represented a buffer enrolled in the NC CREP program that appeared to be ideally situated in the landscape to provide maximum water quality benefits (i.e. upslope pollutant source, no identified drainage ditches, short-circuiting, or deeply incised stream). The primary NO3 − -N source at the site was inorganic fertilizer applied to the adjacent field. Crops rotated on this field included corn, soybeans, peanut, and cotton. The buffers were designed and installed in 1999 by members of the North Carolina Division of Soil and Water, who oversee the NC CREP. The buffer was installed downslope of the agricultural fields and next to an unnamed first-order tributary (Hydrologic Unit 03020102). The tributary flowed into nearby Beech Swamp, which drained into the Fishing Creek watershed. The total length of the combined buffer sections was approximately 304 m (1000 ft). Section 1 had an average width of 60 m (197 ft), while Section 2 had an average width of 45 m (141 ft). The buffer was designed to follow the NRCS recommended three-zone design (USDA-NRCS, 1997). The vegetation at the site was planted based on zone, with areas closest to the stream being dominated by oaks (Quercus spp.), areas midway through the buffer were primarily loblolly pine (Pinus taeda), and the field edge consisted of mainly clover (Trifolium spp.) The NRCS Soil Survey (2006) identified three dominant soil types at the research site: Marlboro fine sandy loam, Bonneau loamy fine sand, and Gritney fine sandy loam. These descriptions were supported with identification from multiple soil cores taken across the site. 2.2. Instrumentation installation Three water table elevation data loggers (Infinities USA, Inc., Port Orange, FL) with built in pressure sensors were installed in the center transect of each zone of the buffer sections (Fig. 1).

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Fig. 1. Position and monitoring schematic of the two buffer sections.

Surficial groundwater monitoring well nests (12 nests in Section 1 and 9 nests in Section 2) were installed in three transects 15 m (50 ft) apart. Each well nest contained a shallow and deep well with maximum depths ranging between 1.5 and 2.3 m (5–7 ft) and 2.7 and 3.6 m (9–12 ft) respectively from the soil surface, and a 0.6 m (2 ft) screened section at the bottom. Section 1 had an additional well nest upslope in each transect to account for its greater width. Four deep aquifer wells (8–11 m deep) were installed at the site to monitor groundwater at a depth believed to be below a restrictive soil layer to assess groundwater mixing between the surficial and deep aquifers in both Sections 1 and 2. Platinum-tipped redox potential probes, constructed as described by Wafer et al. (2004), were installed in the soil next to each of the surficial groundwater monitoring wells in the center transect of Section 1 and the upstream transect of Section 2, to evaluate soil conditions for denitrification potential. The probes were placed at the same depths as the surficial shallow (1.5–2.3 m) and deep (2.7–3.6 m) water quality wells. 2.3. Monitoring and data collection Water table elevation data loggers were used to monitor water table elevation hourly from November 2005 to May 2010, and were downloaded monthly using a HP 48 G+ handheld calculator (Hewlett Packard, Palo Alto, CA). Monthly manual water table elevation readings were collected from the water table wells to calibrate the continuous water table sensors (Solinst® , Georgetown, ON). Water table depths were also recorded in surficial and deep water quality wells prior to groundwater sampling to help provide snapshots of groundwater gradients across the entire buffer. Groundwater samples were collected from January 2005 to May 2010. Prior to sample collection, surficial groundwater wells were purged with a 12 V, low flow submersible pump (Waterra, WSP-12V-5) while an inertial pump (Waterra, Mississauaga, ON) with a foot valve was used for the deep aquifer wells. Flowproportional surface water samples were taken at the upstream

and downstream locations of the buffer with 6712 Portable Teledyne ISCO automated samplers with integrated 730 Bubbler Flow Modules (Teledyne ISCO, Lincoln, NE). All water quality samples were analyzed in the Biological and Agricultural Engineering Environmental Analysis Laboratory for nitrate (NO3 − -N) (EPA Method 353.2) and chloride (Cl− ) (EPA Method 325.2) monthly, while dissolved organic carbon (DOC) (EPA Method 415.1) was analyzed bimonthly for groundwater samples (U.S. EPA, 1983). Monthly sodium (Na+ ) and calcium (Ca2+ ) analyses to determine groundwater signatures began in July 2008 (Standard Method 3111-B) (APHA, 2005). Additionally, soil samples collected at the time of well installation were analyzed for particle size distribution to determine soil hydraulic conductivity at monitoring depths using SPAW 6.0 (NRCS, Pullman, WA). Redox measurements were taken monthly starting in May of 2006 using a KCl-saturated Ag/AgCl reference electrode (Jensen Instruments, Tacoma, WA) and an Accumet AP63 portable pH/mV meter (Fisher Scientific® , Pittsburgh, PA). Measurements were adjusted using a correction factor of 204 mV to account for an average soil temperature of 15 ◦ C (59 ◦ F) and soil pH of 5.2 (Richardson and Vepraskas, 2001). 3. Data analysis 3.1. Hydrology Water table elevations were determined using the site topographic survey, continuously monitored water table elevation data, and monthly manual water table depth measurements. Groundwater flow direction (hydraulic gradients) and soil wetness were assessed using water table data. Groundwater flow direction was modeled using Microsoft Excel 2007® and spreadsheet methods developed by Devlin (2003) along with Golden Surfer 7 mapping software (Golden, CO). Median groundwater velocities were calculated using Darcy’s Law for saturated flow through porous media from water table and soil hydraulic conductivity data. Residence

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time was calculated based on groundwater velocity and buffer width. Wetlands have been shown to have high potential for denitrification (Peterjohn and Correll, 1984; Reddy et al., 1989); therefore buffers were evaluated for wetness using water table proximity to the soil surface and USACE minimum jurisdictional wetland criteria as a benchmark. Minimum jurisdictional wetland hydrology was defined as the water table being within 30 cm of the soil surface consecutively more than 5% (11 days) of the growing season (March 20 to November 6 for Halifax County, NC) in 50% of the years evaluated (USACE, 1987). 3.2. NO3 − -N removal efficiency and NO3 − -N/Cl− ratios Groundwater NO3 − -N removal efficiency was calculated within each of the three transects in each of the buffer locations. NO3 − -N to Cl− ratios were also utilized to determine whether denitrification or dilution was the primary mechanism responsible for groundwater NO3 − -N concentration reductions in the buffer. Lowrance (1992) along with other researchers have used Cl− , a conservative ion (i.e. having minimal plant uptake and not undergoing microbial transformations in soil), to evaluate the impact dilution has on observed groundwater NO3 − -N losses. For example, dilution would be determined a major factor if groundwater NO3 − -N decreased through the buffer toward the stream but the NO3 − -N to Cl− ratio remained constant. 3.3. Measured NO3 − -N mass removal The groundwater NO3 − -N loads for the depths sampled were estimated using Darcy’s Law and the Dupuit–Forchheimer equation (McMahon and Böhlke, 1996; Böhlke et al., 2004; Kennedy et al., 2009) to evaluate the change and/or transformations of groundwater NO3 − -N from the field edge (FE) to the stream edge (SE) within the buffer (Eq. (1)): LoadNO

3

−

-N

=

2 2 − HZone )×K ×T ×W ×C 2.4 × 10−2 × (HZone 3 2

2×L (1)

where LoadNO − -N is the groundwater NO3 − -N flux in the vertical 3 direction for each month (kg N); H is the level of groundwater elevation above datum at specified position (m); K is the hydraulic conductivity at well location (m/h); T is the time (d); C is the influent concentration (mg L−1 ); W is the length of the buffer (m); L is the distance between each groundwater well (m).

Monthly groundwater NO3 − -N loads were calculated using hydraulic conductivities estimated from soil data, hydraulic gradients estimated from hourly monitored water table elevation data, and NO3 − -N concentrations from the 60 cm soil layer at the water quality sampling depths in the surficial aquifer. The LoadNO − -N was 3

reported as kg N m−2 buffer yr−1 through the 60 cm soil layer that corresponded to the water quality sampling depths in each buffer. 3.4. Statistical analysis A statistical analysis was completed to test for significant differences in mean NO3 − -N concentrations throughout the buffer treatment system using SAS PROC MIXED® (SAS Institute, Cary, NC). A log transformation was required to normalize the groundwater NO3 − -N concentrations, the fixed effect was well depth, and the random effects were transects within sections. Redox measurements, Cl− , NO3 − -N/Cl− ratios, DOC, Na+ , and Ca2+ concentrations were considered individual response variables and evaluated with the same procedure as NO3 − -N concentrations. Tukey’s honest significance test was used within each treatment between similar well depths to evaluate changes in groundwater NO3 − -N. Evaluations between the buffer sections and the deeper aquifer water quality constituents were completed using an ANOVA in Proc Mixed (SAS PROC MIXED® , Cary, NC) with NO3 − -N, Cl− , NO3 − -N/Cl− , DOC, Na+ , and Ca2+ concentrations being the individual response variables and the random effects being transects within sections. All statistical tests were considered significant at ˛ = 0.05 level. 4. Results and discussion 4.1. Groundwater NO3 − -N removal within the buffers NO3 − -N removal and influencing factors are summarized in Table 1 for both buffer sections. Groundwater NO3 − -N concentrations at the 1.5 m depth in the wider Section 1 and at the 1.5 m and 3 m depths in the more narrow and downstream Section 2 were significantly reduced through the buffer (˛ = 0.05). Mean groundwater NO3 − -N concentrations at the 1.5 m depth in Section 1 at the field edge (FE) and stream edge (SE) monitoring locations were 4.5 and 1.7 mg L−1 respectively (63% reduction, P = 0.023), while mean groundwater NO3 − -N concentrations in Section 2 at the FE and SE monitoring locations were 12.9 and 1.4 mg L−1 respectively (89% reduction, P < 0.0001). Mean groundwater NO3 − N concentrations in Section 1 from the FE to SE monitoring locations at the 3 m depth decreased from 2.9 to 2.5 mg L−1 respectively

Table 1 Summary of evaluations conducted on the two riparian buffer sections. Section 1 Width Depth of groundwater measurements Mean nitrate at field edge Mean nitrate at stream edge Nitrate reduction efficiency Field edge average elevation Groundwater gradients through buffer Average groundwater velocity Median residence time Water table within 30 cm of soil for 5% of growing season Meets jurisdictional wetland criteria Mean redox near stream Seasonal mean DOC Retention of nitrate/chloride ratios across buffer Measured nitrate removed per year over buffer area (60 cm soil layer)

Section 2 55–60 m

1.5 m 4.5 mg L−1 1.7 mg L−1 63%

40–45 m 3m 2.9 mg L−1 2.5 mg L−1 15%

30.1 m 0.003–0.036 m/m 1.6 cm day−1 3.0 cm day−1 8.7 years 5.8 years Absent in Zones 1 and 3 Present 1 of 5 years in Zone 2

<200 mV 74%

Absent in all zones <200 mV 2.9–12.2 mg L−1 36% 0.003 kg N m−2 yr−1

1.5 m 12.9 mg L−1 1.4 mg L−1 89%

3m 12.8 mg L−1 6.0 mg L−1 54%

28.8 m 0.003–0.010 m/m 1.3 cm day−1 2.8 cm day−1 11.5 years 7.7 years Present 2 of 5 years in Zone 1 Present 4 of 5 years in Zone 2 Absent in Zone 3 Present in Zone 2 <200 mV <200 mV 2.8–14.5 mg L−1 84% 34% 0.06 kg N m−2 yr−1

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a.

Therefore, it was not prudent to make performance comparisons with respect to width. Therefore, landscape setting (i.e. position of each section with respect to the stream and upland source), hydrology (i.e. groundwater flow and direction), and biogeochemistry of the buffer soils (i.e. soil redox) were evaluated to determine how the source of NO3 − -N entered the two buffer sections, and what mechanisms were responsible for the observed reductions within the buffers.

16

Nitrate-N (mg/L)

14 12 10 8 6

A a

B

4

b

b

C C

0 FE (60 m)

MB 1 (50 m) 1.5 m Depth

Nitrate-N (mg/L)

MB 2 (30 m)

SE (1.5 m)

3 m Depth

16 14

A

a b

12 10 B

8

c

6 4 C

2 0 FE (45 m)

MB (30 m) 1.5 m Depth

4.2. Site topography and hydrology

c

2

b.

301

SE (1.5 m)

3 m Depth

−

Fig. 2. Overall mean groundwater NO3 -N concentrations at the 1.5 m and 3 m depths of (a) Section 1 (60 m width) and (b) Section 2 (45 m width). Notes: FE = field edge, MB = mid buffer, SE = stream edge monitoring locations. Values in parentheses are distance of well location from the stream. Error bars represent standard error. Comparisons are shown in each section and well depth independently, as indicated by capital and lower-case letters. Bars with different letters are significantly different at the ˛ = 0.05 level.

(15% reduction, P = 0.784), while NO3 − -N concentrations in the same locations in Section 2 decreased from 12.8 to 6.0 mg L−1 respectively (54% reduction, P = 0.0005). The magnitude of the reductions observed in each buffer section was related to incoming groundwater NO3 − -N source concentrations. Mean groundwater NO3 − -N concentrations entering the FE of Section 1 and 2 were significantly different (P < 0.0001). Groundwater NO3 − -N concentrations entering Section 2 were approximately 3 times higher than concentrations entering Section 1 (Fig. 2). Therefore, since the groundwater concentrations entering Section 1 were low, the observed percent NO3 − -N reductions were also lower than in Section 2. However, mean groundwater NO3 − -N concentrations at the 1.5 m depth discharging to the stream from Sections 1 and 2 were not significantly different (P = 0.193). Mean NO3 − -N concentrations in upstream (1.22 mg L−1 , n = 52) and downstream (5.95 mg L−1 , n = 77) flow weighted samples in the receiving stream were near the same mean concentrations observed in the 3 m well depths at the SE in both Sections 1 and 2. These similarities suggest that the groundwater discharging from the sandy soil layer at the 3 m depth contributed heavily to stream NO3 − -N concentrations, so without this buffer, concentrations in the receiving stream would likely have been much higher. Clearly, the two sections of buffer were receiving groundwater with much different NO3 − -N concentrations, as was also observed at one study site of Angier and McCarty (2008) (discussed earlier).

Buffer elevation and slope often influence water table depth. The average field edge elevation was 29.9 m and 28.9 m in Section 1 and Section 2, respectively. Fig. 3 shows the ground elevation decreased at a relatively constant rate through Section 1, with an average slope of 1.67%. The ground slope in Section 2 decreased substantially from 4% slope between the field edge (FE) and mid buffer (MB) monitoring locations to a 0.3% slope between the MB and stream edge (SE) monitoring locations. Section 2 had a more concave land surface typical of well-defined floodplains. The largest impact of these differences in topography was that the water table was closer to the ground surface in Section 2. Water table elevations and gradients were evaluated to investigate the general groundwater movement across the site and through each buffer section from the FE to SE (Fig. 4). Average water table elevations were higher in Section 1 than in Section 2 during wetter years (2004–2006), while during drier years the water table elevations were more similar (2007–2009). A closer examination of groundwater contour maps generated using groundwater modeling clearly showed the downstream Section 2 had a larger groundwater contributing area from the adjacent field than Section 1, resulting in a greater percentage of groundwater flow toward the lower topographic location (Section 2), regardless of season (Fig. 5). Therefore, more fertilized farmland was draining toward Section 2 than Section 1. This resulted in groundwater with high NO3 − -N concentrations being routed to Section 2, as was observed in the FE groundwater samples. Lower water table elevations in Section 2 indicated local groundwater flowed toward Beech Swamp, less than 2 km downstream. The average difference in water table elevations was smaller in Section 2 (0.13 m) between the FE and MB monitoring locations compared to Section 1 (0.45 m), while the average difference in water table elevations between the MB and SE monitoring locations were similar in Section 2 (0.15 m) and Section 1 (0.16 m). The average water table difference from the FE to the SE was 0.3 m lower in Section 2, suggesting smaller flow gradients between the buffer zones, a result of its lower elevation in the landscape and flat topography. To estimate travel times of the groundwater through the buffers, the hydraulic gradient through these sections was modeled starting in June 2008 using monthly piezometric readings from the water quality wells, a spreadsheet analysis designed by Devlin (2003), and mapping software. Groundwater gradients were calculated based on water table elevation differences from the field edge to the stream edge perpendicularly through the buffer sections (normal to the stream). Section 1 had higher gradients than Section 2, and when coupled with saturated hydraulic conductivity values, it resulted in the fastest groundwater velocity at both the 1.5 m and the 3 m depths (Table 2). The faster groundwater velocity beneath the wider Section 1 buffer resulted in total groundwater travel times that were similar to those calculated in the narrower Section 2 buffer (Table 3). Monthly groundwater residence times varied depending on water table gradients, which were seasonal. However, residence times within the buffer were in the range that was sufficient for

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Fig. 3. Cross section and surficial monitoring wells of (a) Section 1 (60 m width) and (b) Section 2 (45 m width) with average water table elevation.

denitrification throughout the year, since it has been found to occur with residence times as small as 1 month (Dettmann, 2001; Puckett, 2004; Tesoriero et al., 2005). Based on the above analysis, a higher source of groundwater NO3 − -N was delivered from the agricultural fields to the narrower and more downstream buffer Section 2. This explained the higher concentrations observed at the field edge of Section 2. The groundwater passed through both buffer sections before discharging into the stream during most of the monitoring period, and resided within the buffer for periods that have been shown to be adequate for NO3 − -N removal. However, other critical factors such as water table frequency and duration near the soil surface, soil redox,

dissolved organic carbon, and potential dilution were examined to determine whether observed reductions in groundwater NO3 − -N were truly removal that could be linked to microbial denitrification. 4.3. Other indicators related to observed decreased groundwater NO3 − -N concentrations 4.3.1. Riparian buffer relative wetness Riparian areas that have water tables near the soil surface for extended durations can often be classified as riparian wetlands. Wetlands, in general, have been shown to be effective sinks of

Table 2 Groundwater gradient, soil hydraulic conductivity, and mean calculated groundwater velocities at water quality monitoring depths through Sections 1 and 2.

Section 1 (60 m width) Section 2 (45 m width)

Depth (m)

Gradient (m/m)

1.5 3 1.5 3

0.003 –0.036 0.003 –0.010

Saturated hydraulic conductivity (cm h−1 )

Groundwater velocity (cm day−1 ) 1.6 3.0 1.3 2.8

1.9–3.4 3.1–7.2 3.4–4.2 5.1–7.9

Table 3 Median groundwater travel times between each monitoring location in the buffer zones for Sections 1 and 2. Travel time of groundwater

FE to MB (years) MB to SE (years) FE to SE (years)

Section 1

Section 2

1.5 m soil depth

3 m soil depth

1.5 m soil depth

3.7 4.9 8.7

2.4 3.4 5.8

6.1 5.4 11.5

Notes: FE, field edge; MB, mid buffer; SE, stream edge monitoring locations.

3 m soil depth 4.1 3.6 7.7

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303

Fig. 4. Water table elevations for each monitoring location in (a) Section 1 (60 m width) and (b) Section 2 (45 m width) from March 2009 to December 2009. Notes: FE = field edge, MB = mid buffer, SE = stream edge monitoring locations. Values in parentheses are distance of well location from the stream.

Fig. 5. Groundwater contour map of (a) July 2009 (dry period) and (b) January 2009 (wet period). All values are in meters.

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NO3 − -N (Peterjohn and Correll, 1984; Reddy et al., 1989; Humenik et al., 1999; Burchell et al., 2007; Kadlec and Wallace, 2009). A wetland hydrology assessment was completed within the buffer sections to determine which areas could be considered riparian wetlands, in order to assess the potential of these buffers to remove groundwater NO3 − -N. Minimum wetland hydrology criteria (USACE, 1987) were not met within the wider, more upland Section 1 buffer, while the Section 2 buffer area, which was bowl-shaped and located downstream, met the jurisdictional wetland hydrology criteria at the MB monitoring location in 4 out of 5 years. Additionally, the SE monitoring location in Section 2 nearly could be considered a wetland, as it met jurisdictional wetland hydrology in 2 out of 5 years. These results were not surprising as the MB monitoring location in Section 2 displayed characteristics of a riparian floodplain marsh, as the soil surface was often wet, planted pine tree (Pinus taeda) survival was low, and herbaceous wetland vegetation was present. Results further suggested that the water table elevations in Section 2 were not as dramatically influenced by dry periods as in Section 1 (Fig. 4). These results suggested that Section 2 was overall hydrologically better suited for denitrification since the system was wetter more frequently and for longer periods of time.

a.

600 500 400

Soil Redox (mV)

304

300 200 mV 200 100 0 -100 -200 -300 -400 FE MB 1 MB 2 SE FE MB 1 MB 2 SE (60 m) (50 m) (30 m) (1.5m) (60 m) (50 m) (30 m) (1.5m) 1.5 m Depth Highest Reading

b.

3 m Depth

Lowest Reading

Average Reading

600 500

4.3.2. Redox potential and dissolved organic carbon Denitrification occurs in soils with low oxidation/reduction (redox) potentials and high dissolved organic carbon (DOC) concentrations, which were present in both buffer sections. Mean redox values were predominately below 200 mV in Section 1 and Section 2 indicating soil conditions favorable for denitrification (Patrick, 1960; Fielder et al., 2007), so the potential for NO3 − -N reductions in both sections was high for most periods (Fig. 6). Redox values were comparable in Sections 1 and 2. Significant differences between the mean redox measurements in the two buffer sections at the SE 1.5 m depth monitoring location were not found (P = 0.732). However, significant differences were found for the mean redox readings at the 1.5 m depth FE monitoring locations (P = 0.0009). The field-buffer interface in Section 1 was located further upslope from the stream, resulting in mean water table depths that were deeper, and somewhat less conducive for denitrification. Further downslope, the redox measurements taken within the interior of the Section 1 buffer were similar to those in Section 2 and indicated the soil conditions within 45–50 m of the stream were not limiting for denitrification. A DOC assessment was used to evaluate carbon availability in the groundwater and whether it differed between Sections 1 and 2 (Fig. 7). Mean DOC concentrations were not statistically different between the buffer sections at monitoring locations in all zones (˛ = 0.05). However, DOC concentrations varied seasonally through the buffer sections, and the highest concentrations occurred during the summer and winter seasons. Both sections had mean shallow groundwater DOC concentrations in the range of 4–8 mg L−1 , during most of the year, which has been shown to be sufficient to support higher rates of denitrification (Spruill et al., 1997; Knies, 2009). Mean DOC concentrations at the 1.5 m depths were higher than at 3 m depths in both sections at most locations. Lower mean DOC concentrations may have contributed to the higher NO3 − -N concentrations measured at the 3 m depth SE monitoring locations in both sections. Overall, both sections were shown to have the critical components for denitrification to proceed at high rates, including high water tables, low redox readings, and suitable DOC concentrations. However, low incoming groundwater NO3 − -N concentrations entering Section 1 likely limited denitrification rates in that buffer section.

Soil Redox (mV)

400 300 200 mV

200 100 0 -100 -200 -300 -400 FE (45 m)

MB (30 m)

SE (1.5 m)

FE (45 m)

1.5 m Depth Highest Reading

Lowest Reading

MB (30 m)

SE (1.5 m)

3 m Depth Average Reading

Fig. 6. Maximum, minimum, and mean soil redox readings at the 1.5 and 3 m soil depths at differing distances relative to the stream from June 2005 to April 2010 in (a) Section 1 (60 m width, n = 60 total readings from each location) and (b) Section 2 (45 m width, n = 60 total samples from each location). Notes: FE = field edge, MB = mid buffer, SE = stream edge monitoring locations. Values in parentheses are distance of well location from the stream.

4.3.3. Denitrification/dilution assessment Even though the buffers showed a high potential for complete removal of NO3 − -N through denitrification, the potential of dilution was examined. Soil borings indicated a restrictive layer at about 4.6 m (15 ft) below the ground surface that likely separated the surficial and deeper aquifers. However, since restrictive layers can be non-homogeneous, NO3 − -N/Cl− ratios and cation signatures within the groundwater beneath the buffers were used as indicators of dilution. Mean NO3 − -N/Cl− ratios in Section 1 dropped 74% at the 1.5 m depth, and 36% at the 3 m depth from the FE to the SE. Mean NO3 − -N/Cl− ratios in Section 2 decreased from the FE to the SE monitoring locations by 84% at the 1.5 m depth and 34% at the 3 m depth. These values were comparable to observed groundwater NO3 − -N reductions within both buffer sections (Table 1), which supports that dilution within the buffer was minimal (Lowrance, 1992). Schoonover and Willard (2003) found similar mean NO3 − N/Cl− ratios (0.1–0.5) in a NO3 − -N reduction study of a forested riparian buffer. They concluded dilution to be minimal because there were no statistical differences in mean Cl− concentrations

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305

Table 4 Calculated NO3 − -N removal (kg N m−2 yr−1 ) for varying depths and zones in each riparian buffer section. Section

90–150 cm soil layer

240–300 cm soil layer

Total

1 2

0.003 0.02

0 0.04

0.003 0.06

through the buffer, and the ratios decreased at a level similar to observed NO3 − -N. Groundwater dilution through mixing between the surficial and deeper aquifers was unlikely. The chemical analysis of the waters indicated significant differences between the two aquifers. Section 1 had average Ca2+ concentrations of 7 mg L−1 and 12 mg L−1 at the 1.5 m and 3 m field edge monitoring location depths, respectively. Average Ca2+ concentrations in Section 2 at the 1.5 m and 3 m field edge monitoring location depths were 14 mg L−1 and 11 mg L−1 respectively. Average Ca2+ concentrations were 98 mg L−1 in the deeper aquifer (monitored at 8–11 m depth), much higher than in the shallower groundwater in Section 1 and Section 2 (P < 0.0001). These differences in groundwater signatures provided additional strong evidence that interaction between the deeper and surficial aquifers was limited by the restrictive layer encountered during soil borings. 4.4. Estimated mass removal of NO3 − -N through the riparian buffers The hydrologic and biogeochemical factors present in these buffers, coupled with the lack of evidence supporting dilution

a.

10

DOC (mg/L)

8

6

4

2

0 FE (45 m)

MB (30 m) 1.5 m Depth

b.

SE (1.5 m)

3 m Depth

10

from groundwater mixing, suggested that biological activity, presumably denitrification, was the primary mechanism for observed NO3 − -N reduction in both buffer sections. Therefore, the mass removal of NO3 − -N from the buffers was quantified. The overall NO3 − -N mass removal in the 60 cm soil layers surrounding the 1.5 m and 3 m depths were estimated and are shown in Table 4. In Section 1, the wider, upstream buffer that received low concentrations, the rate of NO3 − -N entering the 1.5 m soil layer and the 3 m soil layer at the FE of Section 1 was estimated to be 17 kg N yr−1 and 14 kg N yr−1 , respectively. NO3 − -N discharged to the stream at 5 kg N yr−1 and 15 kg N yr−1 from the 1.5 m soil layer and 3 m soil layer, respectively. NO3 − -N entering the 1.5 m soil layer and the 3 m soil layer at the FE of Section 2, the narrower buffer in the lower topographic location, was 80 kg N yr−1 and 175 kg N yr−1 , respectively; much more heavily loaded than Section 1. NO3 − N discharged into the stream from Section 2 at 5 kg N yr−1 and 25 kg N yr−1 for the 1.5 m soil layer and 3 m soil layer, respectively. NO3 − -N mass removal was seasonally influenced by hydrology and no other factors were observed to significantly limit reductions. The Section 1 buffer reduced groundwater NO3 − -N by 0.003 kg N m−2 yr−1 (76%) in the 1.5 m soil layer, with no change in the 3 m soil layer. The removal in the 1.5 m soil layer of Section 1 was similar to those calculated by Lowrance et al. (1995) (0.002–0.004 kg N m−2 yr−1 ) in buffers also with limited available NO3 − -N. Reduction of NO3 − -N in Section 2 was calculated to be 0.02 kg N m−2 yr−1 (94%) and 0.04 kg N m−2 yr−1 (86%) in the 1.5 m soil layer and 3 m soil layer, respectively. These values were high compared to other buffer studies reviewed, but this buffer had wetland hydrology and received high concentrations of NO3 − -N, elevating its potential for treatment. One study by Nelson et al. (1995) estimated removal rates of 0.012 kg N m−2 yr−1 in a riparian forest area with both somewhat poorly and poorly drained soils, and incoming NO3 − -N concentrations less than those observed in this study. If only median removal rates in the poorly drained portions were considered (where most of the treatment took place), removal rates could be estimated at 0.024 kg N m−2 yr−1 in the Nelson et al. (1995) study, close to the values measured in the 1.5 m depth in Section 2. Additionally, Hoffman et al. (2011) found N retention rates to range between 0.005 and 0.034 kg N m−2 yr−1 in a study of four riparian wetlands, with variability between wetlands also due to nitrogen loading inputs.

8

DOC (mg/L)

5. Conclusions 6

4

2

0 FE (45 m)

MB (30 m) 1.5 m Depth

SE (1.5 m)

3 m Depth

Fig. 7. Average DOC concentrations for (a) Section 1 (60 m width, n = 176) and (b) Section 2 (45 m width, n = 187). Notes: FE = field edge, MB = mid buffer, SE = stream edge monitoring locations. Values in parentheses are distance of well location from the stream. Error bars represent standard error.

The study further supports riparian buffers as an effective BMP for protecting streams from groundwater NO3 − -N originating from agricultural lands. Both sections of the buffer studied showed significant removal of NO3 − -N. The highest removal rates were measured in Section 2, the narrowest buffer, because (1) it received a greater load of the groundwater NO3 − -N from the upland source and (2) it possessed wetland hydrology (based on USACE jurisdiction criteria), which created conditions favorable for NO3 − -N removal through processes such as denitrification. Section 1 also possessed conditions that were favorable for NO3 − -N removal through denitrification. However, because of its landscape position, the source of groundwater NO3 − -N that entered this section was small. The extra width in Section 1 provided no additional water quality benefit and it was likely that this

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buffer could have provided similar reductions even if it had been much narrower. Section 2 appeared to have appropriate width and was in an ideal location to maximize NO3 − -N removal. All results further indicate the importance of site evaluations prior to buffer installations. Many buffer widths and placements are dependent on the landowner and the allowable buffer width supported by conservation programs, as found at this site. Therefore, installed buffer width is rarely a function of meeting NO3 − -N reduction goals for groundwater entering the buffer. Designing riparian buffers relative to groundwater contributing areas, available denitrification enhancing conditions (water table depths close to the soil surface, low redox readings, and available DOC), and entering groundwater NO3 − -N concentrations will improve NO3 − -N removal efficiency within these systems, while preserving valuable land for agricultural practices instead of unnecessarily taking it out of production. To maximize groundwater NO3 − -N removal impacts of buffers enrolled in conservation programs, delineating floodplains, distinguishing ridge lines, and identifying low gradient topographies should be completed prior to land enrollment. This study provided a clear example to conservation buffer program managers that larger buffer width does not imply greater water quality protection. The buffers were positioned in more ideal landscapes than many other sites enrolled in the NC CREP Program (i.e. upslope pollutant source, no identified drainage ditches, short-circuiting, or deeply incised stream). This study confirmed that these areas represented the higher end of potential groundwater NO3 − -N removal within the upper Coastal Plain region of North Carolina. Groundwater NO3 − -N is a major pollutant to streams in eastern North Carolina, so correct riparian buffer placement in the landscape will help ensure that proper hydrology is present to create conditions that will maximize denitrification rates and hence removal. Conservation programs could protect a greater number of stream miles and have a greater impact on overall water quality improvements at the river basin scale if future riparian buffer enrollments are similar to Section 2. Research presented in this paper will be used in outreach efforts to help NC CREP program managers identify ideal widths and locations for future buffer enrollments. Acknowledgements This grant was supported by the NC Department of Environment and Natural Resources – Division of Soil and Water Conservation and the NC Clean Water Management Program to improve the impact of the NC Conservation Reserve Enhancement Program (NC CREP). Special thanks to the staff from these agencies, in particular Natalie Woolard, David Williams, and Charles Bowden. Many thanks to the NCDENR Aquifer Protection Section (Rick Bolich, Evan Kane, and Ray Milosh) who coordinated installation of deep aquifer wells at the site. NCSU staff, undergraduate and graduate students, from the Department of Bio&Ag Engineering provided support for well installation, data collection, and site maintenance, particularly, L.T. Woodlief, Craig Baird, Dale Hyatt, Mike Shaffer, Jamie Blackwell, Jacob Wiseman, and Randall Etheridge. Dan Line was pivotal in his contributions to stream flow and water quality monitoring, and Karen Hall provided assistance in vegetation surveys. We also received advice from Dr. Wendell Gilliam and Dr. Robert Evans on experimental setup. We would also like to thank Dr. Jason Osborne for his assistance with the statistical analysis used in this study. References Angier, J.T., McCarty, G.W., 2008. Variations in base-flow nitrate flux in a first-order stream and riparian zone. J. Am. Water Resour. Assoc. 44 (2), 367–380.

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