OR/18/011 Typical settings

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Stuart, M E. 2018. Review of denitrification potential in groundwater of England. British Geological Survey Internal Report, OR/18/011.

Aquifers

Confined zones

Many studies have confirmed the presence of redox boundaries in confined aquifers similar to Figure 2.2. The boundary occurs some distance down dip from the start of confinement depending on the aquifer flow regime. For the UK these studies are described further in UK aquifers.

In a study of the Triassic sandstone aquifer of Lorraine (LTSA), Celle-Jeanton et al. (2009) [1] found an increase in most of the chemical parameters along the flow path, from unconfined to confined aquifer. This general trend was correlated with the evolution of groundwater residence time and was very close to the conceptual model developed for the East Midlands Aquifer, UK, which showed similar geological and hydrogeological patterns. However, In the LTSA, there was only limited contamination by NO3 with concentrations varying from 1.2 to 14.3 mg/L.

Unconfined zones

An assessment of NO3 concentrations in groundwater in the United States by Burow et al. (2010)[2] indicated that concentrations were highest in shallow groundwater beneath agricultural land use in areas with well-drained soils and oxic geochemical conditions. Concentrations were lowest in deep groundwater where groundwater is reduced, or where groundwater is older and concentrations reflect historically low N application rates. Classification and regression tree analysis was used to identify the relative importance of N inputs, biogeochemical processes, and physical aquifer properties in explaining groundwater NO3 concentrations. The analysis showed that dissolved Fe concentrations explained most of the variation in groundwater NO3 concentration, followed by Mn, Ca, farm N fertilizer inputs, percent well-drained soils and DO. Overall, NO3 concentrations in groundwater were most significantly affected by redox conditions, followed by nonpoint-source N inputs, indicating importantly that redox can overcome N inputs.

Geochemical and isotopic tools were applied at aquifer, transect, and subtransect scales by Hinkle et al. (2007)[3] to provide a framework for understanding sources, transport, and fate of dissolved inorganic N in a sandy aquifer near La Pine, Oregon. NO3 was a common contaminant in shallow ground water in this area, whereas high concentrations of NH4 (up to 39 mg/L as N) were present in deep ground water. N concentrations, N/Cl ratios, tracer-based groundwater ages, N isotope data, and hydraulic gradients indicated that septic tank effluent was the primary source. Nitrogen isotope data, N/Cl and N/C ratios, 3H data, and hydraulic considerations pointed to a natural, sedimentary organic matter source for the NH4. Low recharge rates and flow velocities largely restricted anthropogenic NO3 to isolated plumes within several metres of the water table. Geochemical and isotopic data indicated that denitrification also affected NO3 gradients in the aquifer. Groundwater evolved from oxic to increasingly reduced conditions. Suboxic conditions were achieved after about 15–30 years of transport below the water table. NO3 was denitrified near the oxic/suboxic boundary. Denitrification was characterized at the aquifer scale with a redox boundary approach that captured spatial variability in the distribution of electron donors.

Superficial deposits

Postma et al. (1991) investigated NO3 distribution and reduction processes an unconfined Quaternary sandy aquifer in Denmark. The aquifer was subdivided into an upper 10- to 15-m thick oxic zone that contained O2 and NO3-, and a lower anoxic zone characterized by Fe2+-rich waters. The redox boundary was very sharp, suggesting that reduction processes were fast compared to the rate of downward water transport. Nitrate-contaminated groundwater TDSs were two to four times higher than in groundwater derived from a forested area. The persistence of the high content of TDS in the anoxic zone and the absence of both NO2 and NH4 indicated the downward migration of contaminants and that active NO3 to N2 reduction was taking place. Possible electron donors in the reduced zone of the aquifer were organic matter and small amounts of pyrite. Electron balances based on concentrations of O2, NO3-, SO4- and total inorganic carbon indicated that pyrite was by far the dominant electron donor even though organic matter was much more abundant.

Puckett and Cowdery (2002)[4] used a combination of ground-water modelling, chemical and dissolved gas analyses, and age dating of water to determine the relation between changes in agricultural practices, and NO3 concentrations in ground water of a glacial outwash aquifer in Minnesota. The results revealed a redox zonation throughout the saturated zone with O2 reduction near the water table, NO3 reduction immediately below it, and then a large zone of Fe3+ reduction, with a small area of SO4 reduction and methanogenesis near the end of the transect. Modelling supported the hypothesis that OC was the electron donor. Denitrification rates were small and were limited by the small amounts of OC, 0.01 to 1.45%. In spite of the OC limitation, denitrification was virtually complete because residence time (50–70 years) was sufficient to allow even slow processes to reach completion.

Organic rich recharge

We find mention of increased rates of denitrification under areas where recharge is high in DOM. It is believed that the organotrophic denitrification in groundwater relies mainly on sedimentary, ‘autochthonous’ organic carbon (Ghiorse and Wilson, 1988[5]; Obermann, 1990 [6]). The rate of denitrification is most often related to the amount of dissolved organic carbon (DOC) in porewater or groundwater, or the amount of soluble organic carbon rather than the total amount of solid fraction (Rivett, 2008[7]).

Work by Siemens et al. (2003)[8] under a range of soils, organic-rich plaggic anthrosols and reducing gleyic podzols and eutric gleysols where the water table was high, to determine whether DOC from soils could be a carbon source for organotrophic denitrification. They found that leached DOC contributed negligibly to the denitrification process because the DOC in the soils themselves appeared not to be bioavailable and concentration patterns did not appear to correspond to distribution of denitrification. They concluded that denitrification in the groundwater below was being controlled by the limited translocation of organic carbon to the soils by crop roots. Denitrification is also likely in aquifers affected by infiltration of DOC-rich surface water. See following section

Riparian zones, hyporheic zones and floodplains

Floodplains act as the collection point for groundwater, overland flow and river water (Burt et al., 2002[9]) (Figure 3.1). As a result, the water table is usually close to the surface and the soil and any unsaturated zone will be close to saturation. A number of studies have shown that flow paths are affected by inundation pattern, hydraulic connectivity in the system and the degree of floodplain saturation. On the landscape scale the floodplain is the reactive interface between the upland and the river (Lewandowski and Nützmann, 2010[10]). On the floodplain scale the hyporheic zone is the reactive interface between the aquifer and surface water. In both interfaces the flow paths and flow velocities are of importance for biogeochemical processes. The same processes that occur in the floodplain also occur in the hyporheic zone although higher concentrations in the riverbed sediments indicate more intense processes.

Figure 3.1    Flow paths in the riparian zone (from Stuart and Lapworth (2011)[11].

Heterogeneity

McClain et al. (2003)[12] showed that rates and reactions of biogeochemical processes vary in space and time and these variations are often enhanced at terrestrial-aquatic interfaces. They defined biogeochemical ‘hot spots’ as patches that show disproportionately high reaction rates relative to the surrounding matrix, whereas ‘hot moments’ were short periods of time that exhibit disproportionately high reaction rates relative to longer intervening time periods. Hot spots occur where hydrological flow paths converge with substrates or other flow paths containing complementary or missing reactants. Hot moments occur when episodic hydrological flow paths reactivate and/or mobilize accumulated reactants.

Many previous studies have confined measurements of denitrification and NO3 retention to shallow sediments (<5 cm deep). Stelzer et al. (2011)[13] determined the extent of NO3 processing in deeper sediments of a sand plains stream by measuring denitrification in core sections to a depth of 25 cm and by assessing vertical NO3 profiles, to a depth of 70 cm. Denitrification rates of sediment slurries based on acetylene blocking were higher in shallower core sections. However, core sections deeper than 5 cm accounted for 68% of the mean depth-integrated denitrification rate. Vertical profiles showed that NO3 concentration in shallow ground water was about 10–60% of the NO3 concentration of deep ground water.

Floodplains and riparian zones

Denitrification also occurs in other areas of groundwater where confined conditions exist, such as in riparian zones and floodplains (Hanson et al., 1994[14]; Hill et al., 2000[15]; Lowrance, 1992[16]; Pinay et al., 1998[17]; Watson et al., 2010[18]).

Triska et al. (1993)[19] investigated the terrestrial-aquatic interface beneath the riparian corridor at Little Lost Man Creek, California to assess hydrological and biological control of nutrient fluxes. Subsurface flow paths were defined from the channel toward the riparian zone and also from the riparian zone toward the channel using tracer-injection studies. Solute transport had a rapid channel component and a slow hyporheic flow component. DO concentration in the hyporheic zone ranged from <1.0 to 9.5 mg/L due to permeability variations in bankside sediments and the proportion of stream water in the lateral hyporheic zone. Both nitrification potential and channel exchange decreased with distance from the channel and were absent at sites lacking effective exchange, due to low DO. Denitrification potential was inversely related to channel exchange and was insignificant in channel sediments.

Vidon and Hill (2005)[20] examined the linkages between hydrologic flow paths, patterns of electron donors and acceptors and the importance of denitrification as a NO3 removal mechanism in riparian zones on glacial till and outwash landscapes in Ontario, Canada. NO3-N concentrations in shallow groundwater from adjacent cropland declined from 10–30 mg/L near the field-riparian edge to 1 mg/L in the riparian zones. Chloride data suggested that dilution could not account for this NO3 decline. The riparian zones displayed a well-organized pattern of electron donors and acceptors that resulted from the transport of oxic NO3-rich groundwater to portions of the riparian zones where low DO concentrations and an increase in DOC concentrations were encountered. Work with δ15N and in-situ acetylene injection to piezometers indicated that denitrification was the primary mechanism of NO3 removal in all zones. A shallow water table was not always necessary for efficient NO3 removal by denitrification.

In the Lake Waco Wetland, Texas, NO3 concentrations were reduced by more than 90% in the first 500 m downstream of the inflow, creating a distinct gradient in NO3 concentration along the flow path of water (Scott et al., 2008[21]). The relative importance of sediment denitrification, DNRA and N fixation was assessed along the NO3 concentration gradient in the wetland. Potential denitrification was observed in all months, potential DNRA was observed only in summer months and N fixation was variable. Both sediments and the wetland were NO3 sinks and accounted for 50% of wetland NO3 removal. Sediments were an ammonium source, but the wetland was often a net sink. The importance of DNRA in freshwater sediments appeared to be minor relative to denitrification.

For the Spree River, Germany, Lewandowski and Nützmann (2010)[10] found that the biogeochemical composition of subsurface water showed little temporal variability while spatial heterogeneity was high on the hectometre scale of the study site as well as on the centimetre scale of the bed sediments. Nitrate was eliminated very efficiently by denitrification in the anoxic aquifer of the floodplain while NH4- concentrations increased under anoxic conditions. Ammonium was thought to originate from the mineralization of OM. The redox patchiness of floodplain aquifers favoured NO3 removal, i.e. a temporal and spatial sequence of anoxic and oxic conditions eliminates N.

Hyporheic zones

The hyporheic zone (HZ) is the active zone between the surface stream and groundwater. Figure 3.2 shows a sketch cross-section of the HZ illustrating the common microbially-mediated N transformation processes. Surface waters supply NO3, dissolved organic nitrogen (DON) and DOC to the hyporheic zone. In aerobic regions of the hyporheic zone, DON can be mineralized to NH4, which can be transformed via nitrification to create additional NO3-. The DOC and NO3 also can be retained in the hyporheic zone via microbial assimilation processes. DOC and NO3 entering anaerobic portions of the hyporheic zone can be utilized for denitrification, which produces N2O and N2, which can degas out of the stream system and return to the atmosphere.

Boulton et al. (1998)[22] showed that exchanges of water, nutrients, and organic matter (OM) occur in response to variations in discharge and bed topography and porosity. Upwelling subsurface water supplies stream organisms with nutrients while downwelling stream water provides DO and OM to microbes and invertebrates in the hyporheic zone.

Dynamic gradients exist at all scales in the hyporheic zone and vary temporally (Boulton et al., 1998[22]):

  • Microscale – gradients in redox potential control chemical and microbially mediated nutrient transformations occurring on particle surfaces
  • Stream-reach scale – hydrological exchange and water residence time are reflected in gradients in hyporheic faunal composition, uptake of dissolved organic carbon, and nitrification
  • Catchment scale – the hyporheic corridor concept describes gradients, extending to alluvial aquifers kilometres from the main channel.
Figure 3.2    Simplified longitudinal cross-sectional view of a stream HZ with commonly observed microbially mediated pathways for nitrate transformations (modified from Zarnetske et al., 2011b[23] Labile dissolved organic carbon supply limits hyporheic denitrification. Journal of Geophysical Research: Biogeosciences, Vol. 116, G04036. www.agupubs.onlinelibrary.wiley.com/doi/full/10.1029/2011JG001730.

Across all scales, the functional significance of the HZ relates to its activity and connection with the surface stream. HZs have been identified as potential active areas for denitrification in many studies. Biogeochemical reactions associated with stream nitrogen cycling, such as nitrification and denitrification, can be strongly controlled by water and solute residence times in the HZ.

Denitrification in hyporheic sediments in a small forested stream in Canada was assessed by Duff and Triska (1990)[24]. Samples were obtained along transects perpendicular to the stream at two sites. In-situ denitrification was evident at all locations tested. Denitrifying potentials increased with distance from the stream channel. Both NO3 and DOC decreased over summer in wells at the base of the forested slope. Acetylene-block experiments coupled with the chemistry data suggested that denitrification could modify the chemistry of water during passage through the HZ.

Roley et al. (2012)[25] studied denitrification in the HZ and floodplains of an agriculturally-influenced headwater stream in Indiana, USA. They measured seasonal denitrification rate and NO3 concentrations profiles in stream sediments to examine the potential for N removal in the HZ. Nitrate concentration and denitrification rates declined with depth into the HZ, but denitrification was still measureable to a depth of 20 cm. Floodplain denitrification rates increased over the course of an inundation event particularly where vegetation was present. Deep groundwater tended to be oxic (6.9 mg O2/L) but approached anoxia (0.8 mg O2/L) after passing through shallow, organic carbon-rich sediments, which suggested that the decline in the NO3 concentrations of upwelling ground water was due to denitrification. Collectively the results suggest that there was substantial NO3 removal occurring in deep sediments, below the HZ.

Storey et al. (2004)[26] determined the distribution of nitrogen-transforming processes, and rate-controlling factors within the HZ of an agricultural lowland stream in Ontario, Canada. Physicochemical parameters were measured along a 10 m-long hyporheic flow line between down and upwelling zones. Sediment cores from the streambed surface and from various depths in each zone and water from the corresponding depth was percolated through each core in the laboratory. Denitrification was measured using a 15N-NO3 tracer.

Physicochemical conditions, microbial processes and nitrogen processing distinguished shallow and downwelling zone samples from deeper and upwelling zone samples. Denitrification was highest in surface and downwelling zone cores, despite high O2 levels, probably due to high pore-water NO3 concentrations in these cores and isolation of denitrifying bacteria from O2 in the bulk water by hyporheic biofilms. Denitrification was limited by DO inhibition in the downwelling group, and by NO3 availability in the upwelling group. Strong evidence indicated that DNRA occurred in almost all cores, and outcompeted denitrification. Field patterns and lab experiments indicated that the HZ at this moderately N-rich site was a strong sink for NO3, fitting current theories that predict where HZs are NO3 sinks or sources.

Zarnetske et al. (2011a)[27] used a steady state 15N‐labelled NO3 and Cl tracer to investigate the spatial and temporal conditions controlling denitrification dynamics in the HZ of an upland agricultural stream, Drift Creek, Oregon, USA. They measured solute concentrations (15NO3-, 15N2, as well as NO3-, NH3, DOC, DO, Cl), and hydraulic transport parameters of the reach and along HZ flow paths of an instrumented gravel bar. HZ exchange was observed across the entire bar with flow path lengths up to 4.2 m and median residence times greater than 28.5 h. The HZ transitioned from a net nitrification environment at its head (short residence times) to a net denitrification environment at its tail (long residence times). NO3- increased at short residence times from 0.32 to 0.54 mgN/L until a threshold of 6.9 h and then consistently decreased from 0.54 to 0.03 mgN/L.

Along these same flow paths, declines were seen in DO and DOC. The rates of the DO and DOC removal and net nitrification were greatest during short residence times, while the rate of denitrification was greatest at long residence times. An injection of labile DOC increased the rate of N removal suggesting that denitrification in anaerobic portions of the HZ is limited by labile DOC supply (Zarnetske et al., 2011b). 15NO3- tracing confirmed that a fraction of the NO3- removal was via denitrification across the entire bar HZ. Production of 15N2 across all observed flow paths and residence times indicated that denitrification microsites were present even where nitrification was the net outcome. These findings showed that the HZ was an active nitrogen sink in this system and that the distinction between net nitrification and denitrification in the HZ was a function of residence time and threshold behaviour.

Complex settings

Feast et al. (1998)[28] applied a detailed hydrochemical sampling programme to the River Bure catchment on the Chalk aquifer system of northeast Norfolk to understand the source and fate of NO3. The Chalk aquifer is covered by a variety of superficial deposits including Pleistocene sand and the Lowestoft Till which are up to 30 m thick (Figure 3.3). Modern contaminants of a mainly agricultural origin produced high concentrations of NO3 (>60 mg/L) in the unconfined valley areas, whereas in confined regions the levels of NO3 were low and commonly below detection limits (<0.02 mg/L).

Figure 3.3    Schematic cross-section illustrating processes controlling the distribution of N species in the Chalk of north-east Norfolk (simplified from Feast et al. (1998)[28]).

Samples from within the glacial deposits had δ15N values of 13.7‰ and 9.3‰ while samples from the chalk alone had a narrower range of -2.1‰ to 11.5‰ and from the confined and semi-confined zones +4‰ and +10‰, characteristic of nitrified soil organic nitrogen. There was no apparent relationship between δ15N and NO3 concentration as would be anticipated if in-situ denitrification were controlling NO3 concentrations. However, many Chalk groundwaters had high N2/Ar ratios (39–72) indicating a significant contribution to dissolved N2 from denitrification. Denitrification was believed to be occurring within the overlying glacial deposits, providing a mechanism for naturally improving groundwater quality. δ15N values of low-NO3 groundwaters from the confined zone were isotopically light (-3‰ to +4‰), inconsistent with an origin from denitrification: it is suggested that these waters had a pre-anthropogenic NO3 signature.

Landon et al. (2011)[29] used data from multiple sources to determine interrelations among hydrogeological factors, redox conditions, and temporal and spatial distributions of NO3 in a 2,700-km2 area of the eastern San Joaquin Valley, California, USA. The primary aquifers were a complex sequence of alluvial fan sediments deposited by major tributaries to the San Joaquin River. The main water-bearing units included unconsolidated alluvial-fan deposits and deeper Pleistocene and Pliocene-aged strata, and semi-consolidated rocks derived from fluvial deposits of predominantly andesitic volcanic detritus. Groundwater was predominantly modern, with detectable NO3 and oxic conditions, but some zones were anoxic or mixed. Anoxic conditions were associated with long residence times that occurred near the valley trough and in areas of historical groundwater discharge with shallow depth to water and with interactions of shallow, modern groundwater with soils. NO3 concentrations were significantly lower in anoxic than oxic or mixed redox groundwater, because residence times of anoxic waters exceed the duration of increased pumping and fertilizer use associated with modern agriculture. Effects of redox reactions on NO3 concentrations were relatively minor. Dissolved N2 gas data indicated that denitrification had eliminated >5mg/L NO3–N in about 10% of 39 wells. Increasing NO3 concentrations over time were slightly less prevalent in anoxic than oxic or mixed redox groundwater. Spatial and temporal trends of NO3 were primarily controlled by water and NO3 fluxes of modern land use.

References

  1. ↑ CELLE-JEANTON, H, HUNEAU, F, TRAVI, Y, and EDMUNDS, W M. 2009. Twenty years of groundwater evolution in the Triassic sandstone aquifer of Lorraine: impacts on baseline water quality. Applied Geochemistry, Vol. 24, 1198–1213.
  2. ↑ BUROW, K R, NOLAN, B T, RUPERT, M G, and DUBROVSKY, N M. 2010. Nitrate in groundwater of the United States, 1991–2003. Environmental Science & Technology, Vol. 44, 4988–4997.
  3. ↑ HINKLE, S R, BÖHLKE, J, DUFF, J H, MORGAN, D S, and WEICK, R J. 2007. Aquifer-scale controls on the distribution of nitrate and ammonium in ground water near La Pine, Oregon, USA. Journal of Hydrology, Vol. 333, 486–503.
  4. ↑ PUCKETT, L J, and COWDERY, T K. 2002. Transport and fate of nitrate in a glacial outwash aquifer in relation to ground water age, land use practices, and redox processes. Journal of Environmental Quality, Vol. 31, 782–796.
  5. ↑ GHIORSE, W C, and WILSON, J T. 1988. Microbial ecology of the terrestrial subsurface. Advances in applied microbiology, Vol. 33, 107–172.
  6. ↑ OBERMANN, P. 1990. Significance of anoxic reaction zones in an aquifer in the lower Rhine region. Mitteilungen der Deutschen Bodenkundlichen Gesellschaft, Vol. 60, 249–258.
  7. ↑ RIVETT, M O, BUSS, S R, MORGAN, P, SMITH, J W N, and BEMMENT, C D. 2008. Nitrate attenuation in groundwater: a review of biogeochemical controlling processes. Water Research, Vol. 42, 4215–4232.
  8. ↑ SIEMENS, J, HAAS, M, and KAUPENJOHANN, M. 2003. Dissolved organic matter induced denitrification in subsoils and aquifers? Geoderma, Vol. 113, 253–271.
  9. ↑ BURT, T P, BATES, P D, STEWART, M D, CLAXTON, A J, ANDERSON, M G, and PRICE, D A. 2002. Water table fluctuations within the floodplain of the River Severn, England. Journal of Hydrology, Vol. 262, 1–20.
  10. ↑ 10.0 10.1 LEWANDOWSKI, J, and NÜTZMANN, G. 2010. Nutrient retention and release in a floodplain's aquifer and in the hyporheic zone of a lowland river. Ecological Engineering, Vol. 36, 1156–1166.
  11. ↑ STUART, M E, and LAPWORTH, D J. 2011. A review of processes important in the floodplain setting. British Geological Survey Open Report OR/11/030.
  12. ↑ MCCLAIN, M E, BOYER, E W, DENT, C L, GERGEL, S E, GRIMM, N B, GROFFMAN, P M, HART, S C, HARVEY, J W, JOHNSTON, C A, MAYORGA, E, MCDOWELL, W H, and PINAY, G. 2003. Biogeochemical hot spots and hot moments at the interface of terrestrial and aquatic ecosystems. Ecosystems, Vol. 6, 301–312.
  13. ↑ STELZER, R S, BARTSCH, L A, RICHARDSON, W B, and STRAUSS, E A. 2011. The dark side of the hyporheic zone: depth profiles of nitrogen and its processing in stream sediments. Freshwater Biology, Vol. 56, 2021–2033.
  14. ↑ HANSON, G C, GROFFMAN, P M, and GOLD, A J. 1994. Denitrification in riparian wetlands receiving high and low groundwater nitrate inputs. Journal of Environmental Quality, Vol. 23, 917–922.
  15. ↑ HILL, A R, DEVITO, K J, CAMPAGNOLO, S, and SANMUGADAS, K. 2000. Subsurface denitrification in a forest riparian zone: interactions between hydrology and supplies of nitrate and organic carbon. Biogeochemistry, Vol. 51, 193–223.
  16. ↑ LOWRANCE, R. 1992. Groundwater nitrate and denitrification in a coastal plain riparian forest. Journal of Environmental Quality, Vol. 21, 401–405.
  17. ↑ PINAY, G, RUFFINONI, C, WONDZELL, S, and GAZELLE, F. 1998. Change in groundwater nitrate concentration in a large river floodplain: denitrification, uptake, or mixing? Journal of the North American Benthological Society, 179–189.
  18. ↑ WATSON, T K, KELLOGG, D Q, ADDY, K, GOLD, A J, STOLT, M H, DONOHUE, S W, and GROFFMAN, P M. 2010. Groundwater denitrification capacity of riparian zones in suburban and agricultural watersheds. Journal of the American Water Works Association, Vol. 46, 237–245.
  19. ↑ TRISKA, F J, DUFF, J H, and AVANZINO, R J. 1993. Patterns of hydrological exchange and nutrient transformation in the hyporheic zone of a gravel-bottom stream: examining terrestrial-aquatic linkages. Freshwater Biology, Vol. 29, 259–274.
  20. ↑ VIDON, P, and HILL, A R. 2005. Denitrification and patterns of electron donors and acceptors in eight riparian zones with contrasting hydrogeology. Biogeochemistry, Vol. 71, 259–283.
  21. ↑ SCOTT, J, MCCARTHY, M, GARDNER, W, and DOYLE, R. 2008. Denitrification, dissimilatory nitrate reduction to ammonium, and nitrogen fixation along a nitrate concentration gradient in a created freshwater wetland. Biogeochemistry, Vol. 87, 99–111.
  22. ↑ 22.0 22.1 BOULTON, A J, FINDLAY, S, MARMONIER, P, STANLEY, E H, and VALETT, H M. 1998. The functional significance of the hyporheic zone in streams and rivers. Annual Review of Ecology and Systematics, Vol. 29, 59–81.
  23. ↑ ZARNETSKE, J P, HAGGERTY, R, WONDZELL, S M, and BAKER, M A. 2011b. Labile dissolved organic carbon supply limits hyporheic denitrification. Journal of Geophysical Research: Biogeosciences, Vol. 116, G04036.
  24. ↑ DUFF, J H, and TRISKA, F J. 1990. Denitrifications in sediments from the hyporheic zone adjacent to a small forested stream. Canadian Journal of Fisheries and Aquatic Sciences, Vol. 47, 1140–1147.
  25. ↑ ROLEY, S S, TANK, J L, and WILLIAMS, M A. 2012. Hydrologic connectivity increases denitrification in the hyporheic zone and restored floodplains of an agricultural stream. Journal of Geophysical Research: Biogeosciences, Vol. 117.
  26. ↑ STOREY, R G, WILLIAMS, D D, and FULTHORPE, R R. 2004. Nitrogen processing in the hyporheic zone of a pastoral stream. Biogeochemistry, Vol. 69, 285–313.
  27. ↑ ZARNETSKE, J P, HAGGERTY, R, WONDZELL, S M, and BAKER, M A. 2011a. Dynamics of nitrate production and removal as a function of residence time in the hyporheic zone. Journal of Geophysical Research: Biogeosciences, Vol. 116, G01025.
  28. ↑ 28.0 28.1 FEAST, N A, HISCOCK, K M, DENNIS, P F, and ANDREWS, J N. 1998. Nitrogen isotope hydrochemistry and denitrification within the Chalk aquifer system of north Norfolk, UK. Journal of Hydrology, Vol. 211, 233–252.
  29. ↑ LANDON, M K, GREEN, C T, BELITZ, K, SINGLETON, M J, and ESSER, B K. 2011. Relations of hydrogeologic factors, groundwater reduction-oxidation conditions, and temporal and spatial distributions of nitrate, Central-Eastside San Joaquin Valley, California, USA. Hydrogeology Journal, Vol. 19, 1203–1224.
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