Elsevier

Water Research

Volume 42, Issue 16, October 2008, Pages 4215-4232
Water Research

Review
Nitrate attenuation in groundwater: A review of biogeochemical controlling processes

https://doi.org/10.1016/j.watres.2008.07.020Get rights and content

Abstract

Biogeochemical processes controlling nitrate attenuation in aquifers are critically reviewed. An understanding of the fate of nitrate in groundwater is vital for managing risks associated with nitrate pollution, and to safeguard groundwater supplies and groundwater-dependent surface waters. Denitrification is focused upon as the dominant nitrate attenuation process in groundwater. As denitrifying bacteria are essentially ubiquitous in the subsurface, the critical limiting factors are oxygen and electron donor concentration and availability. Variability in other environmental conditions such as nitrate concentration, nutrient availability, pH, temperature, presence of toxins and microbial acclimation appears to be less important, exerting only secondary influences on denitrification rates. Other nitrate depletion mechanisms such as dissimilatory nitrate reduction to ammonium and assimilation of nitrate into microbial biomass are unlikely to be important in most subsurface settings relative to denitrification. Further research is recommended to improve current understanding on the influence of organic carbon, sulphur and iron electron donors, physical restrictions on microbial activity in dual porosity aquifers, influences of environmental condition (e.g. pH in poorly buffered environments and salinity in coastal or salinized soil settings), co-contaminant influences (particularly the contrasting inhibitory and electron donor influences of pesticides) and improved quantification of denitrification rates in the laboratory and field.

Introduction

Since the 1970s, nitrate (NO3) contamination of groundwater has become a significant environmental problem, with many parts of the world now reporting groundwater nitrate pollution (Burden, 1982, Spalding and Exner, 1993, Beeson and Cook, 2004, European Environment Agency (EEA), 2000, Rao, 2006, Rivett et al., 2007, Roy et al., 2007, Organisation for Economic Co-operation and Development (OECD), 2008). The consequences include long-debated health concerns arising from increased risks of methaemoglobinaemia and cancer (Fan and Steinberg, 1996, WHO, 1999, WHO, 2004, Höring and Chapman, 2004), and environmental impacts such as the eutrophication of surface waters due to excess nutrients (Vitousek et al., 1997, WHO, 1999, Mason, 2002). Diffuse pollution from intensive agriculture since the mid-20th century has largely been blamed for these problems (Foster, 2000). It has been estimated that 70–80% of the nitrate in English surface and groundwaters is derived from agricultural activities (Defra, 2002). However, direct application of nitrogen-based fertilizers to land is not the only source. Discharge from septic tanks and leaking sewers, atmospheric deposition and the spreading of sewage sludge and manure to land can all contribute (Wakida and Lerner, 2005).

The European Union and World Health Organization (WHO) have both set the standard for nitrate in potable water at 11.3 mg nitrogen (N) per litre (50 mg-NO3/l) (Drinking Water Directive 98/83/EC; WHO, 2004). The cost of compliance is already significant (Knapp, 2005). In the UK alone, the cost of treatment to ensure potable water supplies are below 50 mg-NO3/l amounted to £16 million per annum during 1992–1997 (Dalton and Brand-Hardy, 2003) and is predicted to rise to £58 million per annum by 2010 as low-nitrate water for blending becomes scarcer (Defra, 2006).

The European approach to the problem has increasingly recognized the need for integrated protection and management of water resources. The Nitrates Directive (91/676/EEC) requires protection of all natural freshwaters and sets a limit of 50 mg-NO3/l which applies to all groundwater irrespective of its intended use, though it is recognized that much lower N concentrations, possibly around 4.4–8.8 mg-NO3/l (1–2 mg-N/l), may trigger eutrophication in nutrient-poor (oligotrophic) surface waters (James et al., 2005). In more nutrient-rich waters, phosphorus concentrations are more commonly limiting. In addition, the Water Framework Directive (2000/60/EC) requires all groundwater bodies to achieve good status by 2015. The goal of good status includes a test that groundwater concentrations do not exceed statutory limits, including that set by the Nitrates Directive.

The severity of the nitrate problem is illustrated by European Environment Agency (EEA) data on groundwater nitrate concentrations across the EU (EEA, 2007). The proportion of groundwater bodies with mean nitrate concentration >25 mg-NO3/l in 2003 was reported as being ca. 80% in Spain, 50% in the UK, 36% in Germany, 34% in France and 32% in Italy. Notable exceptions were the Scandinavian and Baltic states where less than 3% of groundwater bodies had reported mean nitrate >25 mg-NO3/l. Variably elevated nitrate concentrations are reported elsewhere, for example, Australia (Australian State of the Environment Committee, 2001) and North America (Hudak, 2000, Hinkle et al., 2007).

Mitigation is difficult due to the long-term, diffuse and continuing nature of the problem (Hiscock et al., 2007, Mathias et al., 2007). Options for alleviation are primarily continued implementation of land-use control measures, such as protection zones designed to reduce subsurface nitrate loading (see for example, Silgram et al., 2005; Hiscock et al., 2007; Johnson et al., 2007), and reliance on natural attenuation processes. Our focus is the latter.

We critically review the scientific understanding of biogeochemical processes that control nitrate attenuation in the subsurface and, specifically, consider processes in the unsaturated zone and groundwater environments (soil zone processes such as plant uptake are excluded). Evidence for retardation of nitrate in groundwater has not been identified, though the process has been observed in some soils due to the presence of poorly crystalline materials that carry variable surface charge, and therefore adsorb otherwise inert anions such as nitrate and chloride (Katou et al., 1996, Clay et al., 2004).

Disregarding retardation, processes that cause nitrate mass removal control the attenuation of nitrate. Denitrification is generally recognized as the most significant mass removal process (Korom, 1992, Burt et al., 1999) and is the primary attenuation process evaluated here. We consider the transformation products generated, the roles of the various electron donors and the effects of environmental conditions. The latter includes assessment of the influence of nitrate and oxygen concentration, nutrient availability, pH, temperature, salinity, toxins, pore size and microbial acclimation. The focus is upon denitrification occurring under natural groundwater conditions and as such reference to denitrification-based water treatment studies is typically excluded, particularly where treatment conditions discussed may not occur naturally (e.g. Schnobrich et al., 2007). Where groundwater-specific examples have not been identified, examples of processes in soil are typically used in analogy. We also discuss nitrate depletion mechanisms other than denitrification such as dissimilatory nitrate reduction to ammonium and assimilation of nitrate into microbial biomass.

Our scope is restricted to a critical review of the biogeochemical processes controlling nitrate attenuation. We summarize the science on biogeochemical processes controlling nitrate attenuation potential at any given locality and indicate gaps in the literature worthy of further research. As such the review is distinct, but complementary, to other reviews. These include the well-cited, but now dated, reviews on denitrification microbiology (Knowles, 1982), saturated zone denitrification (Korom, 1992), natural and artificial denitrification by Hiscock et al. (1991) and tracking denitrification (using 15N abundance) by Mariotti (1986). They also include the more recent reviews of non-agricultural sources of groundwater nitrate by Wakida and Lerner (2005), denitrification occurrence across landscapes and waterscapes by Seitzinger et al. (2006), riparian zone nitrogen removal effectiveness by Mayer et al. (2006) and Haycock et al. (1997), methods for measuring denitrification by Groffman et al. (2006) as well as other case study based reviews such as Kinniburgh et al., 1999, Rivett et al., 2007 and Domagalski et al. (2008) who examine field evidence for denitrification within specific aquifer settings.

Section snippets

Electron acceptor context

Bacteria in aquifers obtain energy from the oxidation of organic compounds or inorganic species (e.g. FeS2, Fe2+, Mn2+). Bacteria that use organic carbon as their energy source also tend to use it as a source of cellular carbon (heterotrophism), while those that use inorganic compounds normally use inorganic carbon (mainly from HCO3) for cell construction (autotrophism). Bacteria obtain their energy by mediating chemical reactions typically involving inter-compound electron transfer.

Fig. 1

Organic carbon (heterotrophic denitrification)

Electrons needed for denitrification can originate from the microbial oxidation of organic carbon. Lack of organic carbon to provide energy to heterotrophic micro-organisms (denitrifying bacteria that use organic carbon as the electron donor) is usually identified as the major factor limiting denitrification rates in aquifers (Smith and Duff, 1988, Starr and Gillham, 1993, DeSimone and Howes, 1998, Jacinthe et al., 1998, Devito et al., 2000, Pabich et al., 2001).

Many factors affect the complex

Nitrate concentration

Some workers (e.g. Morris et al., 1988, Smith and Duff, 1988, Korom et al., 2005) have reported that the kinetics of denitrification at concentrations >1 mg-N/l are zero order (i.e. independent of concentration), suggesting that supply of electron donors controls the rate. Excess nitrate concentrations affect the denitrification process by inhibiting the formation of N2 gas and causing the denitrification process to terminate with the formation of N2O (Blackmer and Bremner, 1978). These

Dissimilatory nitrate reduction to ammonium

Dissimilatory nitrate reduction to ammonium (DNRA) is a further anaerobic reduction reaction that can be used by fermentative bacteria (Korom, 1992). This was represented by Robertson et al. (1996) as2H+ + NO3 + 2CH2O  NH4+ + 2CO2 + H2OThe DNRA reaction occurs under much the same conditions as denitrification but is less commonly observed in practice. The partitioning of nitrate between denitrification and DNRA is believed to be controlled by the availability of organic matter: DNRA is the favoured

Conclusions and recommendations

Elevated nitrate concentrations in groundwater may lead to the derogation of precious aquifer resources and the eutrophication of surface waters. Within Europe, these risks are managed through Member States’ implementations of the Nitrates and Water Framework Directives. Understanding of processes controlling the natural attenuation of nitrate, which may lead to risk reduction, is critical to the implementation of these directives.

We have reviewed current understanding of denitrification in the

Acknowledgement

This work was funded by the Environment Agency (for England and Wales) under Environment Agency Science Project SC030155. The views are those of the authors and may not reflect the views or policy of the Environment Agency.

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