OR/18/011 Processes, measurements and indicators
|Stuart, M E. 2018. Review of denitrification potential in groundwater of England. British Geological Survey Internal Report, OR/18/011.|
Nitrogen is an essential nutrient for plant and animal growth, and is cycled in the natural environment in a complex series of biological and chemical processes (Hiscock et al., 1991). In groundwater, denitrification is the dominant nitrate (NO3) attenuation process and becomes active where oxygen is depleted. In aerobic heterotrophic respiration, organic carbon is oxidized preferentially with the electron acceptor that supplies most energy to the microorganisms, namely free oxygen (O2) (Rivett et al., 2008). With an excess of organic carbon, aerobic bacteria consume dissolved oxygen (DO) until it is depleted whereupon reduction of other electron acceptors becomes energetically favourable. Once DO is consumed, facultative anaerobes (bacteria capable of growing with or without O2) use NO3- as an electron acceptor for anaerobic respiration, the denitrification process. The sequence of predominant terminal electron-accepting processes (TEAPs) was set out by Lovley and Chapelle (1995).
Denitrification is the reduction of NO3- normally carried out by anaerobic bacteria to form N species ultimately lost to the atmosphere. This process produces a series of intermediates, including the nitrite ion (NO2-), nitric oxide (NO) and nitrous oxide (N2O), with the final step being the production of gaseous nitrogen (N2). Denitrification is thus a dissimilatory NO3 reduction process with N being lost from the groundwater system. These intermediaries have been used to demonstrate denitrification (Anderson et al., 2014; Bragan et al., 1997; Duff and Triska, 1990; Groffman et al., 1998; Li et al., 2013; Mühlherr and Hiscock, 1997, 1998). Isotopic fractionation has also been used (Granger et al., 2008).
There have been a number of reviews of groundwater denitrification in the UK context. Hiscock et al. (1991) reviewed the necessary environmental conditions for denitrification in groundwater and extended these to artificial denitrification. They stated that most denitrifying bacteria are heterotrophic and are able to utilize a wide range of carbon compounds (sugars, organic acids, amino acids) as electron sources. Nutrient requirements are further discussed by Champ et al. (1979) and Bitton and Gerba (1984). Historical evidence appeared to show that NO3 reduction was not observed at concentrations above 0.2 mg/L (Skerman and MacRae, 1957). This did not take account of more modern concepts of small, protected niches (hotspots) which enable organisms to live in conditions different from the bulk conditions. A biofilm even just a few cells thick can provide enough cover to have an anaerobic layer in an ostensibly aerobic environment. Dependence on pH range and temperature are covered by Gauntlett and Craft (1979). However Rivett et al. (2008) concluded that the critical limiting factors are oxygen tension and electron donor concentration and availability. Variability in other environmental conditions such as NO3 concentration, nutrient availability, pH, temperature, presence of toxins and microbial acclimation appears to be less important, exerting only secondary influences on denitrification rates. Korom (1992) included both denitrification and dissimilatory reduction to ammonia (DNRA) in their review of saturated zone processes. They concluded that natural denitrification can decrease NO3 contamination in modern waters but that it was difficult to predict the rate.
|Electron acceptor||Initial species||Product||Process||ΔG kJ/eq||Relative energy yield||Solution Eh (mV)|
After consumption of NO3-, a further sequence of ions can be used as electron acceptors with decreasing energetic yields in a step-wise process. These include reduction of Mn4+ then Fe3+ to soluble oxidation states (Mn2+ and Fe2+ species) with increase in observed concentrations, reduction of sulphate (SO42-) to S species, and finally the reduction of carbon dioxide to methane (Figure 2.1 and Table 2.1). The presence or absence of these parameters can be used as indicators of low redox and therefore of denitrification potential (McMahon and Chapelle, 2008). This reaction sequence is commonly seen along groundwater flow lines (Edmunds et al., 1982; Edmunds et al., 1984) typically as aquifers become confined.
Most likely denitrifying organisms possess truncated pathways which require synergistic relationships among different denitrifying species to complete reduction to N2 (Jones et al., 2013) and this can be the source of the accumulation of N2O as a greenhouse gas (Müller et al., 2014). Seitzinger et al. (2006) argue that groundwater is an important location for denitrification due to long groundwater residence times, but the uncertainty is large.
Korom (1992) also discuss the denitrification process in the context of addition of NO3 to groundwater where Mn, Fe or SO4 have already been reduced, where autotrophic denitrification may occur using the reduced inorganic compounds as electron donors. They found that groundwater containing Fe2+ did not contain any observable NO3-. Autotrophic denitrification coupled to sulphide or Fe oxidation has been proven for microbiological isolates ((Straub et al., 1996; Weber et al., 2006) but demonstrating this at the field-scale is more difficult. Nitrate reduction by oxidation of pyrite should lead to increased concentrations of SO4. Schwientek et al. (2008) looked for evidence that denitrification could be regulated by pyrite oxidation. A combination of sulphur isotopes coupled with assessment of long (c.100 years) travel times indicated that this was likely to be at a very slow rate. Zhang et al. (2009) showed that NO3 removal from the groundwater below cultivated fields at Oostrum, Netherlands, correlated with SO4 production, and the release of dissolved Fe2+ and pyrite-associated trace metals (e.g. As, Ni, Co and Zn). These results, and the presence of pyrite in the sediment matrix within the NO3 removal zone, indicated that denitrification coupled to pyrite oxidation (autotrophic denitrification) was a major process in the aquifer. A number of modelling studies also indicated that pyrite oxidation was a potential pathway (e.g. Wriedt and Rode, 2006). These processes via Fe or sulphur oxidation are termed chemoautotrophic denitrification (Burgin and Hamilton, 2007). Jahangir et al. (2013) found a positive correlation between groundwater SO4 concentration and denitrification rate and suggest that, due to low DOC in most groundwater environments, denitrification may well be autotrophic. A similar correlation with NH4 was attributed to possible DRNA.
Other N cycle processes
Denitrification is only part of the nitrogen cycle and a number of other processes leading to the immobilisation of N or to the production of similar intermediate species may be operating. These include:
- Dissimilatory nitrate reduction directly to NH4+ (DRNA). DRNA occurs under much the same conditions as denitrification, but is less commonly observed in practice. Compared to NO3-, the resultant NH4+ is a more biologically available and less mobile form of inorganic N (Burgin and Hamilton, 2007). The partitioning of NO3 between denitrification and DNRA is believed to be controlled by the availability of organic matter: DNRA is the favoured process when NO3 (electron acceptor) supplies are limiting and denitrification is favoured when carbon (electron donor) supplies are limiting (Korom, 1992). There are two recognised DRNA pathways, fermentation of organic matter and autotrophic sulphur oxidation, where N2 may be produced (Burgin and Hamilton, 2007).
- Assimilatory nitrate reduction (NO3- => organic N). In the presence of NH4+, this is generally taken up preferentially from water by biomass. It can also be taken up by phreatophytes such as poplars or willows (Rivett et al., 2008).
- Anaerobic ammonium oxidation (anammox) (NO2- + NH4+ => N2). This mainly occurs in soils and in the marine environment and estuaries (Dalsgaard et al., 2005).
- Nitrification (NH4+ => N2O => NO2- and NO3-). Buss et al. (2004) reviewed the attenuation of ammonium, although in the context of landfill leachate contamination, and showed this to be predominantly due to nitrification and to cation exchange predominantly on clay minerals. The presence of N2O as an indicator of denitrification is not necessarily conclusive as it may form from partial nitrification of ammonium (De Groot et al., 1994; Kinniburgh et al., 1999).
- Abiotic processes Burgin and Hamilton (2007) review the relative importance of these processes for the removal of N and conclude that although relatively little is known, they may still be significant in the removal of up to 50% of NO3 loading.
Measurements of denitrification
Groffman et al. (2006) set out the available approaches to demonstrating and quantifying denitrification (Table 2.2). They conclude that this process is very difficult to measure with different problems in various environments and at different scales. They predicted that mass balance and stoichiometric approaches combined with point measurements would be likely to provide the greatest improvements to current understanding. The references provided in Table 2.2 were added from the groundwater literature during compilation of this report. These demonstrate that stable isotopic methods have been the most widely applied. Rivett et al. (2007) also describe some of these methods to confirm NO3 attenuation occurrence. Of the methods shown in Table 2.2, only the mass balance approach appears to be feasible in groundwater without new specialist measurements.
There appear to be very few estimates of denitrification rates in groundwater. Korom (1992) tabulated a range of laboratory denitrification rates for aquifer samples in the range of 0.004 to 1.16 mgN/kg dry sediment/day and for aquifers of <LOD to 3.1 mgN/L/day. They also include half-lives of 1.2 to 2.1 years in sand and gravelly sand from Kölle et al. (1985) and Böttcher et al. (1989). Tesoriero et al. (2000) investigated the rate and mechanisms of NO3 removal in an unconfined sand and gravel aquifer using a series of well nests in the Abbotsford-Sumas aquifer on the west USA-Canadian border. Little or no denitrification was observed in the upland portions of the aquifer but a gradual redox gradient as water moved deeper into the aquifer was observed. A complete loss of NO3 was observed and pyrite oxidation was considered to be the electron source. Denitrification rate estimates were based on mass balance calculations using NO3 and excess N2 coupled with groundwater travel times. Denitrification rates in the deep, upland portions of the aquifer were found to range from <0.01 to 0.14 mM of N per year; rates at the redox gradient along the shallow flow path range from 1.0 to 2.7 mM of N/year.
Zhang et al. (2009) estimated the rate of denitrification in the Oostrum study to be 0.6 mM NO3/year for a 5-m section of depleted aquifer at the top of a well with NO3 concentration of 3 mM in the lower sections.
Jahangir et al. (2013) measured in-situ groundwater denitrification rates in subsoil, at the bedrock interface and in bedrock at two sites in Ireland, grassland and arable, using an in-situ ‘push–pull’ method with 15N-labelled NO3-. Measured groundwater denitrification rates ranged from 1.3 to 469.5 μg N/kg/day Exceptionally high denitrification rates observed at the bedrock interface at the grassland site (470 ± 152 μg N/kg/day) suggest that deep groundwater can serve as substantial hotspots for NO3-N removal. However, denitrification rates at the other locations were low. Denitrification rates were negatively correlated with ambient DO, redox potential, permeability and NO3- (all p values, p<0.01) and positively correlated with SO42- (p<0.05). A higher mean N2O/(N2O+N2) ratios at an arable site (0.28) compared to a grassland site (0.10) revealed that the arable site had higher potential to indirect N2O emissions.
|Acetylene-based methods||Inhibit the reduction of N2O to N2||N2O||Bragan et al. (1997); Groffman et al. (1999); Mühlherr and Hiscock (1997, 1998); Weymann et al. (2008)|
|15N tracers||Addition of radio-labelled NO3||Isotope fractionation, or dilution, 15N mass balance, direct measurement of 15N-labelled gases||Christensen et al. (1990). Few recent references for groundwater|
|Direct N2 quantification||Direct estimation from soil||N2||Vogel et al. (1981); Wilson et al. (1990)|
|N2:Ar quantification||Estimate excess N2||Needs precise MS measurements||Blicher-Mathiesen et al. (1998)|
|Mass balance approaches||Estimate by difference||Measure all other fluxes-generally whole catchment||Böhlke et al. (2002); Lindsey et al. (2003); Tesoriero et al. (2000)|
|Stoichiometric approaches||Uses constant C:N:P in organic matter to estimate N losses||Problems with surface water due to changes in C & P||Works for marine environment|
|Stable isotopes to measure isotopic fractionation||Use δ15N and δ18O (plus S and C isotopes)||Mass spectrometry||Aravena and Robertson (1998); Böhlke and Denver (1995); Böhlke et al. (2002); Böttcher et al. (1990); Chen and MacQuarrie (2005); Fukada et al. (2003); Kellman and Hillaire-Marcel (1998|
|In-situ gradients||Use environmental tracers as age indicators||CFCs, SFsub>6, 3H, 14C||Böhlke et al. (2002); Cook and Böhlke (2000)|
|Molecular approaches||Study microbial community||Functional genes in denitrification pathway-DNA probes, PCR||Barrett et al. (2013); Wakelin et al. (2011)|
Keuskamp et al. (2012) attempted to model the extent of denitrification and N2O production at the European scale using an improved version of the IMAGE model. They assumed a first order decay process with a range of half-lives for NO3, based on lithology, texture and effective porosity. These ranged from 1 year for silici-clastic medium fine consolidated rocks (pyrite-bearing), 2 years for alluvial materials to 5 years for unconsolidated coarse sediments, based on the data from Korom, (2002). In all river basins modelled deep groundwater NO3 outflow was much less than the outflow from shallow groundwater. Modelled N2O production was very heterogeneous, with large regions having no groundwater and therefore short travel times or no deep groundwater, and therefore no denitrification or N2O production expected.
Other potential indicators
Redox sequence ions
It is anticipated that there will be a sequence of redox changes as water migrates from upland recharge areas to lowland discharge areas under confined conditions. Champ et al. (1979) identified 3 zones O2–NO3, Fe-Mn, sulphide. Hiscock et al. (1991) showed change in redox potential is often accompanied by a sequential reduction in dissolved groundwater species which is sited as proof of denitrification.
This reaction sequence is commonly seen along groundwater flow lines (Edmunds et al., 1982; Edmunds et al., 1984) typically as aquifers become confined. Water at recharge is generally saturated with DO at the partial pressure of the atmosphere (10–12 mg/L depending upon barometric conditions). Passing through the soil and the unsaturated zone some of this O2 will react as a result of microbiological processes and oxidation-reduction reactions. However, almost all water reaching the water table still contains several mg/L O2. Geochemical and microbial reactions progressively remove the O2 along flow lines. Once all the O2 has reacted an abrupt change of water chemistry takes place (redox boundary). Down-gradient of the redox boundary, denitrification occurs and it is likely that Fe2+ concentrations will increase. Sulphate reduction and the production of sulphide (H2S as S2- in solution) may also occur at greater depths (Figure 2.2).
Denitrification can also be important in shallow groundwater where recharge contains an elevated loading of organic carbon.(Smith et al., 1991; Spalding and Parrott, 1994; Zarnetske et al., 2011b). In their review of floodplain processes, Stuart and Lapworth (2011) tabulated a set of criteria characterising the redox zones (Table 2.3) based on earlier work. These used a series of indicators including the electron acceptors O2, NO3 and SO4, intermediates NO2 and N2O and the solid phase acceptors Mn4 and Fe3, indicated by the presence of dissolved Mn and Fe, and eventually methane. The values shown in Table 2.3 were applied to landfill leachate plumes and some concentrations, particularly for Fe, are high.
|Respiration||Energy yielding process|
|O2 mg/L||NO3 mgN/L||NO2 mgN/L||N2O µg/L||NH4 mgN/L||Mn mg/L||Fe mg/L||S mg/L||SO4 mgS/L||CH4 mg/L|
Water quality criteria (mg/L)
|Reference for criteria|
|Oxic||O2 reduction||≥0.5||-||<0.05||<0.1||Seitzinger et al. (2006); Tiedje (1988)|
|Suboxic||<0.5||>0.5||<0.05||<0.1||Further definition not possible|
|Mn4+ reduction||<0.5||<0.5||≥0.05||<0.1||Chapelle and McMahon (1991); Chapelle et al. (1995); Christensen et al. (2000); Elliot et al. (1999); Murphy and Schramke (1998); Plummer et al. (1990)|
|Fe3+/SO42- reduction||<0.5||<0.5||-||≥0.1||≤0.5||Chapelle et al. (2002)|
|Mixed||-||-||-||Criteria for more than one process are met|
McMahon and Chapelle (2008) evaluated redox conditions in 15 principal aquifers across the USA. They used a series of indicators including the electron acceptors DO, NO3 and SO4, and the solid phase acceptors Mn4 and Fe3, indicated by the presence of dissolved Mn2+ and Fe2+ (Table 2.4). Parameters included were chosen because they were relatively inexpensive and easy to measure, and most of them were commonly measured in regional water quality assessments. They recommended that other redox indicators, such as NH4, H2S, CH4, and H2, should be measured whenever possible. An internally consistent set of threshold criteria was developed and applied to the aquifers using data from the NWQA Program. These were then related to both natural (As) and anthropogenic (NO3 and VOCs) contaminant distribution. The redox conditions explained many of the observed water quality trends at the regional scale. Identification of zones of redox heterogeneity were also important for assessing the fate and transport of contaminants. This approach is similar to that of Lyngkilde and Christensen (1992) and Bjerg et al. (1995) above but uses lower concentrations.
Zhang et al. (2009) studied NO3 removal from a sandy aquifer below cultivated fields and forested areas at Oostrum in the Netherlands. Nitrate loss correlated with SO4 production, and increase in dissolved Fe2+ and pyrite-associated trace metals (e.g. As, Ni, Co and Zn) in a zone between 10 and 20 m deep. These results indicated that denitrification coupled to pyrite oxidation is a major process in the aquifer. Significant NO3 loss coupled to SO4 production was further confirmed by comparing historical estimates of regional SO4 and NO3 loadings to age-dated groundwater SO4 and NO3 concentrations, for the period 1950–2000. This highlights a warning against always anticipating SO4 removal as redox decreases.
Lee et al. (2008) characterised the redox conditions in an arsenic-affected aquifer in the Lanyang Plain, Taiwan. Discriminant analysis was adopted to delineate three redox zones (oxidative, transitional and reductive zones) in different aquifers and yielded over 90% agreement with groundwater quality data. According to the DA results, the groundwater of the Lanyang plain was classified as Zone 1 (oxidizing zone) if DO ≥ 1 mg/L or NO3 ≥0.35 mg/L, Zone 2 (transitional zone), if DO was between 0.3 and 1 mg/L or NO3 < 0.35 mg/and Zone 3 (reducing zone) if DO <0.3 mg/L or HS-≥ 0.07 mg/L.
Predisposing settings approach
Smith et al. (2009) proposed a classification scheme for pollutant attenuation at the groundwater-surface water interface/mixing zone. They assumed that a variety of factors influence the rate and magnitude of natural attenuation processes in the subsurface. These included contaminant-specific properties, such as structure, composition, sorption properties and recalcitrance, and environmental properties including sediment geochemistry, hydrochemistry, transport velocity and residence time in reactive zones. The primary properties included were sediment thickness and permeability (as a measure of relative residence time within the GW–SW interface), groundwater baseflow index as a measure of groundwater-stream connectivity, and sediment geochemistry using cation exchange capacity, the fraction of organic carbon (foc) and total inorganic carbon (TIC) to represent the retardation potential.
They postulated that in the case of denitrification the foc of the sediments may be used as a proxy measure of denitrification potential, on the assumption that denitrifying bacteria are ubiquitous in the environment and that foc is a reasonable indicator of redox conditions (Rivett et al., 2008). Where the foc value is high, anaerobic conditions conducive to denitrification are more likely to dominate. Thresholds were proposed as 3% organic carbon content along a riparian flow line based on work by Dahl et al. (2007) and Hoffman et al. (2000). The work was focussed on data collected under river baseflow conditions and analysis using national scale data and was related to surface water aquifer conditions.
Absence of predicted N
Denitrification can be inferred from the absence of groundwater concentrations which would be predicted from the nitrogen source term. Inputs could include N applications to arable land and urban areas, and aerial deposition. NVZS.
An example of this approach was produced by Stuart et al. (2017). Figure 2.3 uses data compiled for the NVZ designation process to highlight areas where N applications are high but groundwater concentrations do not reflect this and the overall risk is not high enough for NVZ designation. Areas where this may be due to a long travel time from the surface have been excluded by defining a travel time of <10 years.
N in rainfall
Nitrogen species in rainfall recharge has been estimated for the individual areas covered by the BGS Baseline Survey. These reports indicate that evaporated rainfall, concentrations multiplied by 3, is likely to contain between 0.75 and 5.5 mg/L of N as NO3 and NH4. Assuming this is all oxidised to NO3, this would represent between 3.2 and 23 5 mg/L in recharge. Lowest concentrations were measured in North Yorkshire and Wessex, and in other sites in the west and southwest and, and the highest in Derbyshire and the North Downs. Areas in central England tended to have about 2–3 mg/L N in evaporated rainfall. Highest concentrations of NO3 were estimated for the North Downs, Derbyshire Carboniferous Limestone and sites up the east coast. Highest concentrations of NH4 were also estimated in East Anglia.
Clearly, this simple pattern does not allow for a number of processes that could intercept N including runoff and plant up-take.
RoTAP (2012) have also modelled nitrogen concentrations in rainfall. Figure 2.4 shows both aerosol and volume weighted concentrations of NO3 in precipitation in 2008. Note precipitation units in µeq/L. The lower limit (11 µeq/L) corresponds to 0.7 mg/L NO3 (0.22 mg/L as N) and the upper (32 µeq/L) to 1.98 mg/L (0.49 mg/L NO3 as N). Multiplying these by 3 as above indicates areas of Wales and Western Scotland to be receiving <0.66 mg/L NO3-N and East Anglia to be receiving rainfall with >1.5 NO3-N.
A similar RoTAP map for reduced N species indicates a similar distribution for particulate NH3 with masses about 1.5 times higher, but with higher masses extending to London and to the extreme southwest. The conversion into concentrations in precipitation was not estimated but it can be assumed that the amount deposited could be 1.5 times that of NO3. Gaseous concentrations however have a different distribution.
This could suggest N deposition in rainfall to be between <1.7 mg/L and >3.8 mg/L or assuming this all to be oxidised to NO3 as <7.6 to >17 mg/L. This broadly agrees with estimates made in the baseline reports.
Keuskamp et al. (2012) describe a spatial model for simulating the fate of N in both soil and groundwater with the aim of predicting both N leaching and N2O emissions. This used an improved version of the global IMAGE model (Bouwman et al., 2006) which takes spatial heterogeneity at the European scale into account.
This model uses a 1×1 km grid and computes a steady state annual water balance at the surface.
- Runoff is partitioned into three depth classes, surface runoff in the top 1 metre, shallow groundwater in the next 5 metres draining to local surface water at short distances and deep groundwater below 6 metres and 50 m thick draining to large rivers at greater distances.
- Travel time distribution calculated by using lithological classes and the time distribution described by Meinardi (1994) assuming that the deep system is fed by a vertically, uniformly draining shallow system.
- N leaching factors were based on the MITERRA model (Velthof et al., 2009) and required N inputs per land use type, soil texture, rooting depth, soil temperature and soil organic carbon.
- Denitrification was calculated using a hole-in-the-pipe model and assuming a first order decay process. In this model, the rate of NO flux from soils depends on both the ‘flow through the pipe’ — the rates of nitrification and denitrification — and the size of the ‘holes in the pipe,’ which were determined by environmental conditions such as soil moisture and temperature (Firestone and Davidson, 1989). The more N ‘flowing through the pipe,’ the higher NO fluxes will be for a given set of environmental conditions.
The model predicted that travel times were 0.5 to 10 years in shallow groundwater and mainly between 50 and 200 years for deep systems. At the European scale only what appear to be areas of the Chalk were plotted on the distribution map for the UK. None of the large rivers systems for which a detailed nitrogen budget has been reported are in the UK. Calculated NO3 delivery to the Thames, Severn and Trent was close to reported values whereas it was greater for the Mersey.
In a national assessment, the groundwater nitrate reduction for catchments without monitoring data, was estimated by a statistical model (Blicher-Mathiesen et al., 2014). The N reduction was estimated as the fraction of aerobic/anoxic water flux above and the anaerobic water flux beneath the redox cline in 1 km2 grids. The depth to the redox cline was defined from the changes in the sediment colour from 12 000 national boreholes or from geological description. The estimated N reduction was calibrated to the N reduction calculated for 56 large (>50 km2) monitored catchments. The results showed areas over 80% of the loading was removed between the root zone and the marine receptor.
Linear discriminant analysis (LDA) for predicting groundwater redox status was applied to a major dairy farming region in New Zealand (Wilson et al., 2018). Data cases were developed by assigning a redox status to samples derived from a regional groundwater quality database. This was based on NO3 <0.5 mg/L, Mn >0.5 mg/L and DO <1.0 mg/L. The majority of groundwater samples were from inland alluvial aquifers.
Pre-existing regional-scale geospatial databases were used as training variables for the discriminant functions. Variables included elevation, slope, drainage, land use and water depth. The models predicted 23% of the region as being reducing at shallow depths (<15 m), and 37% at medium depths (15–75 m). Predictions were made at a sub-regional level to determine whether improvements could be made with discriminant functions trained by local data. The results indicated that any gains in predictive success were offset by loss of confidence in the predictions due to the reduction in the number of samples used.
The regional scale model predictions indicated that subsurface reducing conditions predominated at low elevations on the coastal plains where poorly drained soils are widespread. Additional indicators for subsurface denitrification are a high carbon content of the soil, a shallow water table, and low-permeability clastic sediments.
A map for spatially variable nitrate reduction in groundwater covering six countries in the Baltic Sea Basin, Denmark, Sweden, Finland, Lithuania, Poland and Germany was developed (Højberg et al., 2017) designed to provide an independent estimate of the nitrogen reduction in groundwater. The total area discharging to the Baltic Sea is about 1 720 270 km2 spanning large climatic and land use variations. Mapping was based on review of national data and studies including process-based models, source apportionment models and hydrological and drainage maps. Depending on availability, different approaches were used for the countries ranging from national modelling for Denmark and Sweden to expert judgement for Poland.
The review revealed large variations in the hydrogeochemical conditions important for transport and degradation of nitrogen in groundwater. This included the hydrogeology, the reducing conditions of the subsurface, and the fraction of water transported by drainage systems bypassing the reducing subsurface environments. Significant variations in groundwater reduction between the countries and within most of the countries were found, indicating that strategies for nitrogen regulation and mitigation measures may be optimised, if variation in the natural reduction of nitrate is considered.
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