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TECHNICAL REPORTS

Ground Water Quality

Persistent Elevated Nitrate in a Riparian Zone Aquifer

Dep. of Earth and Environmental Sciences, Univ. of Waterloo, 200 University Ave. W., Waterloo, ON, Canada

^* Corresponding author (wroberts{at}sciborg.uwaterloo.ca).

Received for publication February 23, 2007.

	ABSTRACT

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Summary and Conclusions REFERENCES

 
Streamside vegetated buffer strips (riparian zones) are often assumed to be zones of ground water nitrate (NO₃^–) attenuation. At a site in southwestern Ontario (Zorra site), detailed monitoring revealed that elevated NO₃^––N (4–93 mg L^–1) persisted throughout a 100-m-wide riparian floodplain. Typical of riparian zones, the site has a soil zone of recent river alluvium that is organic carbon (OC) rich (36 ± 16 g kg^–1). This material is underlain by an older glacial outwash aquifer with a much lower OC content (2.3 ± 2.5 g kg^–1). Examination of NO₃^–, Cl^–, SO₄^2–, and dissolved organic carbon (DOC) concentrations; N/Cl ratios; and NO₃^– isotopic composition (¹⁵N and ¹⁸O) provides evidence of four distinct NO₃^– source zones within the riparian environment. Denitrification occurs but is incomplete and is restricted to a narrow interval located within 0.5 m of the alluvium–aquifer contact and to one zone (poultry manure compost zone) where elevated DOC persists from the source. In older ground water close to the river discharge point, denitrification remains insufficient to substantially deplete NO₃^–. Overall, denitrification related specifically to the riparian environment is limited at this site. The persistence of NO₃^– in the aquifer at this site is a consequence of its Pleistocene age and resulting low OC content, in contrast to recent fluvial sediments in modern agricultural terrain, which, even if permeable, usually have zones enriched in labile OC. Thus, sediment age and origin are additional factors that should be considered when assessing the potential for riparian zone denitrification.

Abbreviations: DO, dissolved oxygen • DOC, dissolved organic carbon • K, hydraulic conductivity • OC, organic carbon

	INTRODUCTION

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Summary and Conclusions REFERENCES

 
VEGETATED riparian buffer strips separating agricultural lands from surface water courses can stimulate natural attenuation of nitrate (NO₃^–) by providing a carbon-rich environment where plant uptake and denitrification Eq. [1] can occur (Korom, 1992; Hill, 1996):

[1]

Studies have reported higher rates of NO₃^– removal in organic carbon (OC)-rich riparian soil zones, compared with underlying C-horizon material (Lowrance, 1992; Flite et al., 2001), and in patches of C-horizon material that have locally higher OC because of the presence of decaying root material (Gold et al., 1998; Jacinthe et al., 1998). However, because of weathering reactions that result in the formation of clay minerals, soil zones normally have hydraulic conductivity (K) values that are orders of magnitude less than that of the sand and gravel aquifers that are most prone to nitrate contamination. Thus, although nitrate consumption rates may be high in carbon-rich soil zones, at sites where thicker aquifers are present, ground water may approach stream discharge points via deeper, more permeable flowpaths and may have less opportunity to interact with surficial carbon-rich layers. Nonetheless, a number of studies have reported substantial NO₃^– attenuation in riparian zones with permeable units present that extend into adjacent upland areas (Blicher-Mathieson and Hoffman, 1999; Mengis et al., 1999; Devito et al., 2000; Puckett et al., 2002; Bohlke et al., 2002). However, in each of these cases the aquifer flowpaths entered OC-rich surface layers because the aquifers pinched out or because upward hydraulic gradients were present. In the latter study, biogenic pyrite associated with surficial peat deposits was also an important electron donor for denitrification. Additionally, upward flow (artesian) conditions cause upwelling of older-aged ground water, which further complicates the interpretation of apparent NO₃^– trends (Bohlke and Denver, 1995; Bohlke et al., 2002; Puckett et al., 2002). Vidon and Hill (2004a), Hill et al. (2004), and Kellogg et al. (2005) reported active denitrification within riparian zone aquifers but attributed this to the presence of zones of OC enrichment within the aquifers. Enrichment of OC became more pronounced in areas near the streams.

This study investigates a field site that offers additional insight into the nature of aquifer–soil zone interaction in riparian zones because it has several favorable characteristics: (i) stratigraphy is simple, consisting only of a soil zone of recent river alluvium and an underlying Pleistocene-aged sand and gravel aquifer with elevated NO₃^– concentrations from nearby agricultural activity; (ii) the contact between the two units is sharp and occurs at a consistent depth across the study site; and (iii) except near the river discharge point, upward ground water flow does not occur, and ground water flow is primarily horizontal. Additionally, the ground water residence time within the riparian zone is substantial (3 yr), and a very detailed, local-scale, monitoring network of 70 monitoring points has been established.

Although direct upward ground water flow from the aquifer into the overlying alluvium does not occur at this site, several possible mechanisms exist for downward transport of soil zone carbon into the aquifer. These include molecular diffusion, hydrodynamic dispersion along the contact zone, periodic downward advection during recharge events, and, if plant roots penetrate into the upper aquifer, decay of root material and root exudates (e.g., Gold et al., 1998; Jacinthe et al., 1998). Our objective was to assess the extent to which an OC-rich riparian soil zone could influence the persistence of nitrate in an underlying aquifer in such a hydrogeologic setting.

	Materials and Methods

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Summary and Conclusions REFERENCES

 
Site Description
The study site (Zorra site) is located in southwestern Ontario (43°00' N, 81°00'W) (Fig. 1 ) in a flat lying agricultural area that is dominated by corn production. The recommended fertilizer N-application rate for corn production in this region ranges from 100 to 200 kg N ha^–1 yr^–1 (OMAF, 2003). The upgradient farm field, in addition to corn production, includes a 2-ha yard area where, beginning in 2001, poultry manure has been composted for resale. Poultry manure has a high N content (e.g., 36 g kg^–1) (Liebhardt et al., 1979) and was thus expected to contribute to N loading at the site. A third-order stream (i.e., Thames River), presumed to be the regional ground water discharge point, is present and is bounded by a 100-m wide riparian zone (Fig. 1), which is the focus of this study. The riparian zone coincides with the floodplain of the river and is naturally vegetated with native grasses, willow shrubs, and some larger deciduous trees (Fig. 2 ). The water table within the riparian zone varies seasonally from ground surface to about 1-m depth (Robertson et al., 2007). Topography suggests that the ground water flow system extends about 1 km upgradient (northwestward) and encounters areas primarily under corn cultivation. During 2005 the floodplain was inundated on at least three occasions; thus, frequent flooding contributes to recent sedimentation in the soil/alluvium layer. The aquifer is much older Pleistocene-aged sediment associated with a regional glacial spillway complex (Cowan, 1971; Chapman and Putnam, 1984). Initial site investigations in 2004 revealed the presence of elevated ground water NO₃^––N concentrations of up to several tens of milligrams per liter, typical of many shallow aquifers in this intensively cultivated region of southern Ontario (e.g., Rudolph et al., 1998). Because of the presence of elevated NO₃^– and the accessibility of the shallow aquifer, the site was selected for testing of a permeable reactive subsurface barrier design that used wood-particle media to promote denitrification (Robertson et al., 2007). In 2004 a reactive barrier 8 x 4 m in surface area and 0.6 m in thickness was installed into the shallow aquifer at 0.9 to 1.5 m depth at a location near the upgradient edge of the riparian zone (Fig. 1).

View larger version (20K): [in this window] [in a new window]   Fig. 1. Zorra riparian zone, showing Section B-B', and nitrate-reactive barrier with associated plume of nitrate depleted ground water.

View larger version (21K): [in this window] [in a new window]   Fig. 2. Zorra riparian zone Section B-B', stratigraphy, monitoring locations, and ground water flow directions.

To assess ground water discharging to the river, six minipiezometers were advanced by percussion into the stream bed to depths of up to 1.5 m, and five seepage meters similar to those described by Lee (1977) were installed into the stream bed along Section B-B' (Fig. 2). The streambed sediment was primarily sand; thus, the seepage meters could be easily pressed into the sediment to depths of 5 to 10 cm to provide hydraulic isolation from surface flow.

Sampling
Sampling in this study was limited primarily to a single "snapshot" undertaken during the peak of the growing season (13–29 June 2005) when biological activity and carbon exchange between the alluvium and the shallow aquifer should have been relatively active. Ground water samples were collected from all of the piezometer bundles on Section B-B' using a peristaltic pump and were filtered (0.45 µm) in-line before atmospheric exposure. Samples for analysis of Cl^–, NO₃^–, SO₄^2–, and dissolved organic carbon (DOC) were collected untreated in 20-cm³ polyethylene sample bottles. Filtered samples (0.3–1 L) were collected separately in 1-L polyethylene bottles for analysis of NO₃^– isotopic composition (¹⁵N and ¹⁸O).

Ground water discharging to the riverbed was sampled on 29 June 2005 using the minipiezometers and seepage meters. Sampling occurred after the seepage meters had been in place for at least 24 h by attaching evacuated plastic bags to the discharge tubing and allowing the bags to fill over a 1- to 2-h period. Filtered samples were retained for measurement of water quality parameters. Samples were kept chilled (4°C) until analysis. Nests 70, 71, 73, and 75 were subsequently sampled (October 2006) for dissolved oxygen (DO) content using a colorimetric field test kit (kit K-7512; CHEMetrics, Calverton, VA).

Sediment Characterization
Continuous undisturbed sediment cores were collected to 3 m depth at each of the eight bundle piezometer locations on Section B-B' (Nests 38, 70–76) (Fig. 2). Cores were obtained in 5-cm-diameter aluminum core tubes advanced into the sediment by percussion. Core tubes were cut length-wise using a saw blade and were stratigraphically logged, and 26 5-cm-long subsamples were selected for grain size and OC analysis. Organic carbon was analyzed at the University of Guelph, ON, Soil and Nutrient Laboratory, by measurement of mass loss on ignition (Heiri et al., 2001), which provided a detection limit of 0.1 g kg^–1 OC. A factor of 0.58 was used to convert organic matter mass loss to an equivalent OC mass. Grain size distribution was determined using sieves. Hydraulic conductivity was estimated from the grain size data using the Hazen method (Freeze and Cherry, 1979).

Vadose Zone Pore water
To better characterize possible N sources affecting the riparian zone, vadose zone pore water from the poultry manure composting yard was sampled on 8 Sept. 2005. Using a soil auger, sediment was retrieved to the water table depth (2.3 m) at a location near the edge of the composting yard (Site 82) (Fig. 1), where yard runoff water periodically pooled during precipitation events. Thus, vadose zone pore water at this location was expected to be dominated by leachate from the composting operation. The sediment was silty sand loam to 0.7-m depth, which was underlain by C-horizon material of fine-coarse sand similar to the underlying aquifer. The sediment was separated into 0.3-m depth increments, retained in plastic bags, and returned to the laboratory for pore water extraction. The extraction procedure involved mixing the sediment samples (3 kg each), with 1 L of deionized water in a plastic bucket, decanting the supernatant solution, and then filtering a 20-mL sample for Cl^–, NO₃^–, NH₄⁺, and DOC analysis and a separate 150- to 500-mL sample for analysis of NO₃^––¹⁵N and ¹⁸O. Although this extraction procedure presumably led to solute dilution, it did not effect NO₃^– isotopic composition or characteristic solute ratios.

Chemical Analyses
Nitrate, Cl^–, and SO₄^2– were analyzed in the Earth Sciences Department, University of Waterloo, ON, by ion chromatography using a Dionex ICS-90 (Dionex, Sunnyvale, CA). This provided a detection limit of <0.5 mg L^–1 for each of these parameters. Dissolved organic carbon was analyzed in the Earth Sciences Department using a Dohrman DC-190 total carbon analyzer (Dohrman, Santa Clara, CA), which provided a detection limit of 1 mg L^–1. Ammonium was analyzed at the Soil and Nutrient Laboratory, University of Guelph, ON, using a colorimetric technique with a Technicon TRAACS-800 autoanalyzer (Technicon Instruments, Tarrytown, NY), which provide a detection limit of 0.02 mg L^–1 as N. Nitrate for isotopic characterization was captured by precipitating as AgNO₃ and converted to N₂ gas for ¹⁵N analysis following the procedures of Kendall and Grim (1990) and Silva et al. (2000). Analysis of ¹⁸O was conducted by combustion of the AgNO₃ salt with baked carbon to CO₂ (Silva et al., 2000). Isotopic analyses were completed at the University of Waterloo, Environmental Isotope Laboratory using a VG PRISM Series II mass spectrometer (GV Instruments, Manchester, UK) for NO₃^––¹⁸O and a Carlo Erba 1108 CNOS Elemental Analyser (Thermo Fisher Scientific, Milan, Italy) interfaced to a Fisons Instruments Isochrom-EA mass spectrometer (GV Instruments, Manchester, UK) for NO₃^––¹⁵N. Results are reported in the standard -notation relative to the reference standards of air for ¹⁵N and Vienna Standard Mean Ocean Water for ¹⁸O. Laboratory duplicate variability averaged 0.5 permil (n = 12) for ¹⁵N and 0.3 permil (n = 5) for ¹⁸O.

The significance of chemical differences between zones was assessed using a standard Student's t test.

	Results

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Sediment Properties
The soil layer consists of organic-rich sandy silt river alluvium, which is underlain at 1 to 1.5 m depth by a sand and gravel aquifer (Fig. 2). The contact between the two units is abrupt and remains remarkably planar across the site (Fig. 2). The alluvium (K 1.2 ± 0.5 x 10^–3 cm s^–1; n = 8) (Fig. 3 ) is 30 times less permeable than the aquifer (K 3.4 ± 2.7 x 10^–2 cm s^–1; n = 12) but has much higher OC content (36 ± 16 g kg^–1; n = 8) (Fig. 3) compared with the aquifer (2.3 ± 2.5 g kg ^–1; n = 12). Although the aquifer is heterogenous and contains some finer silty zones, these remain OC depleted (e.g., bottom of cores 70 and 72, OC 0.8–1.5 g kg^–1) (Fig. 3). The total thickness of the aquifer at the site is unknown.

View larger version (21K): [in this window] [in a new window]   Fig. 3. Depth profiles of sediment (a) organic carbon (OC) content and (b) hydraulic conductivity (K) values determined from core samples from piezometer nests 70, 72, 74, and 76.

Ground Water Flow
Previous investigations associated with the NO₃^– barrier, which was installed to 1.5 m depth at the edge of the riparian zone (Robertson et al., 2007), showed that after 11 mo of operation, a plume of NO₃^––depleted ground water extended southeastward from the barrier for a distance of 25 m toward the river (Fig. 1). This revealed the direction of ground water flow and established a mean annual ground water velocity in the shallow aquifer of 28 m yr^–1. This was important evidence because in permeable floodplains ground water flow paths do not always approach surface water discharge points orthogonally but can deflect in the downstream direction (Woessner, 2000; Vidon and Hill, 2004b). Additionally, the horizontal hydraulic gradient along Section B-B' was flat lying and thus prone to measurement error and varied seasonally (0.0046–0.0003 m m^–1), making velocity calculations using the Darcy equation difficult. The velocity estimate of 28 m yr^–1 indicates that ground water resides for 3 yr within the 100-m wide riparian zone. No detectable vertical hydraulic gradients were present within the aquifer (<1-cm head change over 3 m depth), but on several occasions after rainfall events, higher head values were observed in the alluvium layer compared with the underlying aquifer, indicating downward flow at those times.

NO₃ Distribution and Source
Composition of the seepage meter water, river water, vadose zone pore water, and the ground water are given in Tables 1 through 4 and Fig. 4 through 6. Four distinct ground water zones can be identified containing NO₃^– from different sources, which are referred to as the Cornfield, Local, Compost, and Old zones (Fig. 4–6).

View larger version (36K): [in this window] [in a new window]   Fig. 4. Section B-B' (a) NO3––N, (b) Cl–, and (c) SO42– concentrations, June 2005.

View larger version (30K): [in this window] [in a new window]   Fig. 6. Section B-B' (a) NO3––15N and (b) NO3––18O, June 2005.

View larger version (30K): [in this window] [in a new window]   Fig. 5. Section B-B' (a) N/Cl ratios, (b) dissolved organic carbon (DOC), June 2005, and (c) dissolved oxygen (DO), October, 2006.

Local Zone
The shallow aquifer at Nests 71 and 72 has lower NO₃^––N (4–12 mg L^–1) (Fig. 4) with somewhat enriched ¹⁵N (+18 to +22 permil) (Fig. 6) and much higher Cl^– (48–64 mg L^–1) (Fig. 4) compared with the upgradient Cornfield zone. This could reflect temporal changes in fertilizer use in the Cornfield source zone, but more likely this is ground water that has been recharged locally within the floodplain. The high Cl^– values, above that expected from evapotranspirative enrichment of precipitation, indicate that this is predominantly recharge from river water. Flooding to a height of up to 0.6 m above ground surface has been observed at the site. The Cl^– values are similar to the river water values during flood events (45 mg L^–1) (Table 2). At Nests 71 and 72, the Local zone occupies the top of the aquifer at 1 to 2 m depth, but farther downgradient (Nests 73–76) the Local zone is less distinct and is possibly absent.

Old Zone
In the downgradient area of the riparian zone, NO₃^––N concentrations in the aquifer (19 ± 3 mg L^–1) (Table 4, Fig. 4) and NO₃^– isotopic signatures (¹⁵N 22 ± 1 permil; ¹⁸O 7.4 ± 0.9 permil) (Table 4, Fig. 6) remained remarkably uniform, in contrast to the upgradient areas, which are more variable. This is interpreted as representing older ground water recharged through the upgradient farm field before its usage as a composting yard or from other locations farther upgradient.

	Discussion

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Compost Zone Isotopic Signature
Nitrate ¹⁵N values in the vadose zone leachates and the Compost zone ground water (+20 to +41.4 permil) (Table 3, Fig. 6) are higher than values measured in many other agricultural studies (e.g., Kreitler, 1979; Komor and Anderson, 1993; Wassenaar, 1995). Nitrogen from synthetic fertilizers normally has ¹⁵N in the range of –5 to +5 permil, whereas manure N typically has ¹⁵N in the range of +8 to +21 permil (Kreitler, 1979; Heaton, 1986; Korom, 1992; Exner and Spalding, 1994). The highly enriched values observed here are indicative of a manure source but clearly have been influenced by additional processes leading to isotopic enrichment. The two most likely processes are NH₃ volatilization (Flipse and Bonner, 1985; Wassenaar, 1995; Karr et al., 2001, Karr et al., 2002) and denitrification (Kreitler, 1979; Böttcher et al., 1990; Korom, 1992), both of which lead to isotopic enrichment of the residual N.

Several previous studies have reported enriched ¹⁵N values in manure ponds and associated ground water, in the range of +21 to +43 permil (similar to this study), and in each case volatilization or denitrification was assumed to be the cause of the enrichment (Komor and Anderson, 1993; Karr et al., 2001; Karr et al., 2002). Volatilization is likely important at the Zorra site due to the mechanical mixing and the extended period of curing (several months) associated with the composting operation. Vadose zone NO₃^––¹⁵N values are variable, with highly enriched values (+32 to +39 permil) present in the shallow zone underlain by less enriched values (+23 to +26 permil) (Table 3). This may reflect seasonal variations in NH₃ volatilization because it has been shown that ¹⁵N in manure lagoons can be more enriched in the summer when maximum volatilization occurs (Karr et al., 2001). An excellent correlation exists between NO₃^––¹⁸O and NO₃^––¹⁵N in the vadose zone pore water (Fig. 7b ), but the slope (0.2) is lower than expected for denitrification (0.5–0.7) (Böttcher et al., 1990; Aravena and Robertson, 1998; Mengis et al., 1999; Devito et al., 2000). Volatilization of NH₃ during curing would also be accompanied by a decrease in compost moisture due to volatilization of water. Evaporative loss of water would increase the ¹⁸O of H₂O, concurrent with an increase in ¹⁵N of NH₃. Nitrification of NH₃ to produce NO₃^– incorporates one oxygen atom from atmospheric O₂ and two oxygen atoms from water (Hollocher, 1981; Anderson and Hooper, 1983; Yoshinari and Wahlen, 1985). Thus, the correlation between NO₃^––¹⁵N and NO₃^––¹⁸O in the vadose zone pore water could reflect the combined effects of volatilization on the NH₃ and H₂O precursors. The elevated DOC values in the vadose zone leachate (21–102 mg L^–1) (Table 3) support the possibility of denitrification also occurring, but the lack of NO₃^––¹⁸O and ¹⁵N enrichment together at the expected slope of 0.5 to 0.7 (Fig. 7b, unsaturated zone) shows that denitrification is unlikely to be solely responsible for the enrichment. Thus, both volatilization and denitrification may contribute to the isotopic imprint of the compost sourced NO₃^–.

View larger version (21K): [in this window] [in a new window]   Fig. 7. Comparison of (a) NO3––N concentration vs. NO3––15N and (b) NO3––15N vs. NO3––d18O in unsaturated zone pore water, ground water, river water, and seepage meter samples.

The Local zone ground water has enriched NO₃^––¹⁵N (+18 to +22 permil) and NO₃^––¹⁸O (+8 to +9 permil) compared with the river water source (Fig. 6 and 7) and thus has likely been affected by denitrification. Presumably, denitrification has occurred during downward migration through the alluvium layer or during lateral migration within the upper aquifer, possibly enhanced by the occasional presence of rootlets in this zone (e.g., Gold et al., 1998; Blazejewski et al., 2005). However, NO₃^––N removal is incomplete, and low concentrations (4–9 mg L^–1) (Fig. 4) remain.

In the Old zone farther downgradient, NO₃^––N concentrations (19 ± 3 mg L^–1) (Table 4) and ¹⁵N values (+21 to +24 permil) (Fig. 6) remain remarkably uniform; thus, there is little evidence of active denitrification. Several Old zone and Compost zone monitoring points located within 0.5 m of the alluvium–aquifer contact have modestly lower NO₃^––N (6–13 mg L^–1; Nests 73, 74, 38, and 75) (Fig. 4). However, enrichment of ¹⁵N is only 2 to 5 permil compared with the underlying ground water (Fig. 6). It is of interest to examine the amount of NO₃^– removal that this amount of isotopic enrichment indicates. Consider the 2.0-m deep piezometer at Nest 75, which lies just below the alluvium/aquifer contact and has NO₃^––N of 8 mg L^–1 (Fig. 4), with a NO₃^––¹⁵N value of +24 permil (Fig. 6). This represents enrichment of about 3 permil compared with the underlying ground water. Using the range of isotopic enrichment factors observed during denitrification in other ground water studies (–5 to –30 permil) (Lehmann et al., 2003), the Rayleigh distillation equation (Lehmann et al., 2003) predicts that a 3-permil enrichment would result from NO₃^––N depletion of 10 to 45% (1–7 mg L^–1).

The minipiezometers had NO₃^––¹⁵N values of +22 ± 1 permil (Table 4, Fig. 6), and four of the five seepage meters had NO₃^––¹⁵N values of +19.8 to +23.5 permil (Table 1), consistent with discharging Old zone ground water without substantial NO₃^– loss by denitrification. One of the seepage meters (SM5) (Table 1) and the shallowest monitoring point in the streamside nest (i.e., Nest 76) had relatively low NO₃^––N values (5–8 mg L^–1) (Fig. 4), which could be interpreted as indicating partial denitrification. However, NO₃^––¹⁵N values in these samples were isotopically light (+13 and +14.6 permil) (Table 1, Fig. 6) and were similar to the river water from the same sampling period (+14.0 permil, 23 June 2005) (Table 2). This indicates that these samples represent streambed interflow within a hyporheic zone (e.g., Woessner, 2000) rather than discharging ground water.

Limitations to Denitrification
The riparian aquifer had detectable DO throughout, but values were low (0.2–0.7 mg L^–1) (Fig. 5c) and were probably not high enough to inhibit denitrification. Thus, rapid denitrification is limited by an absence of suitably labile electron donors within the riparian flowsystem. Although the aquifer OC (2.3 ± 2.5 g kg ^–1) (Fig. 3) is not capable of supporting high rates of denitrification, slower denitrification is still possible. Considering the previous estimate of maximum NO₃^––N loss in Piezometer 75 at the 2.0-m depth (7 mg L^–1 or 1.5 mg N kg^–1 of sediment, assuming porosity of 0.35 and bulk density of 1.7 g mL^–1, typical of quartz sand), the 3-yr transit time of ground water through the riparian zone to this location indicates a reaction rate of up to 1.4 µg N kg^–1 d^–1. This should be considered a maximum rate for the riparian aquifer. Higher rates are possible within the Compost zone where DOC is elevated. However, the amount of depletion in this zone is difficult to estimate because NO₃^– concentrations are more variable. Kellogg et al. (2005) measured similarly low NO₃^– reaction rates (<1.5 µg N kg^–1 d^–1) at two sites in a glacial outwash aquifer and an alluvial riparian aquifer in Rhode Island, but these low values were accompanied by much higher rates (30–120 µg N kg^–1 d^–1) particularly within 10 m of the stream discharge points. Increased reaction rates near the streams were attributed to the presence of zones of OC enrichment. In this study, Nest 76, located within 1 m of the river bank, exhibited modestly elevated OC (5.8–9.0 g kg^–1) (Fig. 3) in the shallow aquifer zone, but this material was apparently incapable of supporting high rates of denitrification or was too limited in extent to substantially deplete NO₃^– entering the river. The next closest monitoring location with OC data (Nest 74, located 35 m from the river) showed no indication of OC enrichment in the aquifer (Fig. 3). Trudell et al. (1986) and Starr and Gillham (1993) measured low rates of denitrification (0.4–7 µg N kg^–1 d^–1) in an unconfined glacio-lacustrine sand aquifer in Ontario (Hillman Creek site). Although low, these rates were capable of substantially depleting high NO₃^––N concentrations (>10 mg L^–1) within the deeper aquifer where the ground water was older. At the Zorra site, however, such low rates would not be capable of substantially depleting NO₃^– because of the high initial concentrations and the limited ground water residence time of only 3 yr within the riparian zone.

In a survey of a number of riparian zones in southern Ontario, Vidon and Hill (2004a) noted rapid depletion of nitrate in five of six sites with aquifers present. This was attributed to the presence of labile OC within the aquifer sediments because they were of recent fluvial origin (Hill et al., 2004). At the Zorra site, however, zones enriched in labile OC capable of supporting high rates of denitrification were apparently absent because the flow system remained entirely within glacial outwash deposits and did not encounter recent fluvial sediments. Although the glacial outwash was of fluvial origin, it was deposited under landscape conditions that were likely much different from today, with little soil development and sparse vegetative cover. Thus, sediments deposited in this environment were likely to be OC poor compared with recent fluvial material, and any labile OC that may have originally been present has had substantial opportunity to be leached out during the 10,000-yr period that these sediments have been in place. A possible reason why recent fluvial sediments were absent at depth in the Zorra floodplain is that coarse gravel deposits were locally present, containing abundant cobbles and boulders. These may have restricted normal stream geomorphological processes. In landscapes such as southern Ontario, however, it is a common occurrence for modern river courses to follow the same topographic depressions left by the glacial spillway channels (Chapman and Putnam, 1984). Thus, conditions similar to the Zorra site could be common in these areas.

In some Pleistocene aquifers with low OC content, nitrate reduction can still occur by using other electron donors, such as pyrite. Autotrophic denitrification using pyrite normally causes a stoichiometric increase in SO₄^2– concentrations (Kolle et al., 1985; Postma et al., 1991; Robertson et al., 1996; Aravena and Robertson, 1998; Tesoriero et al., 2000). However, SO₄^2– concentrations remained relatively consistent within each of the ground water zones (Fig. 4), which argues against the occurrence of this attenuation mechanism. Iron can also act as an electron donor (Postma et al., 1991), but Fe²⁺ concentrations in the ground water at the site were too low (<0.5 mg L^–1) (Robertson et al., 2007) to substantially deplete NO₃^–.

	Summary and Conclusions

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Summary and Conclusions REFERENCES

 
At this site, direct upward ground water flow from the aquifer into the surficial OC-rich alluvium does not occur. Possible secondary mechanisms of carbon exchange between the two units (e.g., molecular diffusion, hydrodynamic dispersion, periodic downward flow during recharge events, root penetration into the aquifer, etc.) were ineffective in substantially depleting NO₃^––N concentrations of 20 mg L^–1 reaching the river. This example demonstrates that when ground water remains within permeable but OC-poor, stratigraphic units when approaching streamside discharge points, interaction with surface carbon-rich zones and denitrification can be limited. In this case the riparian aquifer was OC poor because it was of Pleistocene age, whereas recent fluvial sediments in modern agricultural terrain often have zones enriched in labile OC, which seem to be the key to NO₃^– attenuation. Although Pleistocene aquifers normally contain some OC (e.g., Blazejewski et al., 2005), which may support denitrification (Trudell et al., 1986; Starr and Gillham, 1993; Puckett, 2004; Kellogg et al., 2005), because of the age of these sediments, the most labile material may be depleted. Reaction rates may be insufficient to substantially remove NO₃^– considering the often limited residence time of ground water within riparian zones. Thus, when assessing the potential for NO₃^– attenuation in riparian zones, evaluation of sediment age and origin may be important additional factors that should be considered.

Manure storage areas have the potential to increase NO₃^– loading to adjacent ground water flow systems. At this site, however, the presence of elevated DOC within the Compost zone ground water (14–68 mg L^–1) leaves open the possibility that additional slow NO₃^– depletion may occur before this ground water discharges to the river.

Although our study was limited to a single site, in landscapes such as southern Ontario it is a common occurrence for modern river courses to follow the same topographic depressions left by the glacial spillway channels (Chapman and Putnam, 1984). Thus, conditions similar to the Zorra site could be widespread in these areas.

	NOTES

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Summary and Conclusions REFERENCES

 
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	REFERENCES

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Summary and Conclusions REFERENCES

 

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