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Assessment of excess N2 and groundwater N2O emission factors of nitrate-contaminated aquifers in northern Germany D. Weymann, R. Well, H. Flessa, C. von der Heide, M. Deurer, K. Meyer, C. Konrad, and W. Walther Soil Science of Temperate and Boreal Ecosystems, Büsgen-Institute, University of Göttingen, Büsgenweg 2, 37077 Göttingen, Germany Inst. for Soil Science, Univ. of Hannover, Herrenhäuser Str. 2, 30419 Hannover, Germany HortResearch, Tennent Drive, Palmerston North, 4474 New Zealand Geries Ingenieure, Büro für Standorterkundung, Kirchberg 12, 37130 Gleichen, Germany Inst. for Groundwater Management, Dresden Univ. of Technology, 01062 Dresden, Germany Received: 12 February 2008 – Accepted: 23 February 2008 – Published: 1 April 2008 Correspondence to: R. Well (rwell@gwdg.de) Published by Copernicus Publications on behalf of the European Geosciences Union.


Introduction
Denitrification is considered the most important reaction for nitrate (NO − 3 ) remediation in aquifers.This process occurs in O 2 depleted layers with available electron donors (Ross, 1995;B öttcher et al., 1990).Especially in agricultural areas with high N inputs via fertilizers considerable NO − 3 reduction is possible (B öttcher et al., 1985).Dinitrogen (N 2 ) is the final product of this process.Thus the quantification of groundwater N 2 arising from denitrification (excess N 2 ) can facilitate the reconstruction of historical N inputs, because NO − 3 loss is derivable from the sum of denitrification products (B öhlke and Denver, 1995).Generally, the concentration of excess N 2 produced by denitrification in groundwater is estimated by comparing the measured concentrations of Ar and N 2 with those expected from atmospheric equilibrium, assuming that the noble gas Ar is a stable component (Blicher-Mathiesen et al., 1998;B öhlke, 2002;Dunkle et al., 1993;Mookherji et al. 2003).However, measuring of excess N 2 is complicated by variations of recharge temperatures and entrapment of air bubbles near the groundwater surface which leads to varying background concentrations of dissolved N 2 in groundwater due to contact of the water with atmospheric air (B öhlke, 2002).Furthermore, N 2 can be lost by degassing (Blicher-Mathiesen et al., 1998).Another aspect of denitrification are potential accumulation and emission of the greenhouse gas nitrous oxide (N 2 O) which represents an obligate intermediate of the process.In contrast to direct agricultural N 2 O emissions arising at the sites of agricultural production, e.g.soils, indirect emissions from ground and surface waters are associated with nitrogen leaching and runoff to adjacent systems (Well et al., 2005a;Nevison, 2000).The knowledge of these indirect emissions is limited because few studies have tried to relate subsurface N 2 O concentrations to N leaching from soils (Clough et al., 2005) and investigations of N 2 O in deeper aquifers are rare (Ronen et al., 1988;McMahon et al., 2000;Hiscock et al., 2002).
In the aquifers of unconsolidated pleistocene deposits covering large areas in the northern part of central Europe, agricultural NO − 3 contamination often coincides with Introduction

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Full reducing conditions (Walther, 1999), suggesting that this region might be susceptible for relatively high N 2 O fluxes from deeper groundwater.However, until now there have been no systematic investigations of N 2 O dynamics in these aquifers.N 2 O emissions from groundwater were thought to comprise a significant fraction of total agricultural N 2 O emissions (IPCC, 1997), but recent studies show in agreement that their significance is presumably lower (McMahon et al., 2000;Hiscock et al., 2003;H öll et al., 2005;Reay et al., 2005;Well et al., 2005a;Sawamoto et al., 2005).Consequently, the nitrous oxide emission factor from aquifers and agricultural drainage water was corrected downwards from 0.015 to 0.0025 by the IPCC in 2006, taking the data of Hiscock et al. (2002Hiscock et al. ( , 2003)), Reay et al. (2004Reay et al. ( , 2005) ) and Sawamoto et al. (2005) as a basis.
Principally, the N 2 O emission factor of a system is defined by the ratio between N 2 O emission and N input (IPCC, 1997).However, the IPCC factor characterizing indirect emissions from aquifers and drainage ditches (EF5-g) had been derived from the ratio between dissolved N 2 O und NO − 3 concentrations observed in a small number of studies, because input and emission data had not been available.Consequently, there are uncertainties in the estimate of EF5-g because both NO − 3 and N 2 O are subject to change during subsurface transport (Dobbie and Smith, 2003).Furthermore, determination of N 2 O fluxes from aquifers is connected with experimental difficulties: N 2 O as an intermediate product from denitrification is permanently influenced by different enzyme kinetics of various denitrifying communities and groundwater N 2 O concentration is the net result of simultaneous production and reduction reactions (Well et a. 2005b In this study we measured excess N 2 and N 2 O in groundwater of 4 nitratecontaminated, denitrifying aquifers in Northwest Germany in order (1) to estimate initial NO − 3 that enter the groundwater surface, (2) to assess potential indirect emissions of N 2 O, and (3) to compare existing concepts of groundwater N 2 O emission factors.

Study sites
Investigations were conducted in the aquifers of 4 drinking water catchments (Fuhrberg, G öttingen, Th ülsfelde and Sulingen) located in Northwest Germany, Lower Saxony.These aquifers consist of pleistocene sand and pleistocene gravel and are characterized by NO − 3 contamination that results from intensive agricultural N inputs via fertilizers.In all aquifers, NO − 3 concentrations in the deeper groundwater are substantially lower compared to the shallow groundwater.In previous studies, denitrification was identified as the natural process for reduction of groundwater NO − 3 concentrations in Fuhrberg (K ölle et al., 1985; B öttcher et al., 1990), Th ülsfelde (P ätsch, 2006;Walther et al., 2001), and Sulingen (Konrad, 2007).General properties of the aquifers are summarized in Table 1.

Sampling and laboratory analyses
Groundwater samples (3 or 4 replications per depth, respectively) were collected during single (Sulingen, G öttingen) or repeated sampling events (Th ülsfelde) or 4 times within one year (Fuhrberg), respectively, from groundwater monitoring wells allowing collection of samples from defined depths (Table 1).The Fuhrberg site was equipped with multilevel sampling wells (B öttcher et al., 1985) with a depth resolution of 0.2 m in the first 2 m of the groundwater and 1.0 m for the rest.Samples were collected using a peristaltic pump (Masterflex, COLE-PARMER, Vernon Hills, USA).Because negative Introduction

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Full pressure in the suction tubing might cause partial outgassing of the water sample during pumping, a low suction rate of approximately 50 ml min −1 was used to minimize this effect.In Fuhrberg, additional samples were collected from taps at the pump outlets of drinking water wells which delivered raw water to the waterworks.The other sites were equipped with regular monitoring wells consisting of PVC-pipes (diameter between 1.5 and 4 ) with filter elements of one or two m length.Here, samples were collected with a submersible pump (GRUNDFOS MP1, Bjerringbro, Denmark), which prevents outgassing because the water samples are at a positive pressure during pumping.
From one of these monitoring wells, replicate groundwater samples were collected using both pump types in order to estimate potential outgassing using the peristaltic pump.Differences between the treatments were non-significant, which proves that outgassing was negligible.For both pump types, groundwater was collected from the outlet through a 4 mm ID PVC tubing by placing its end to the bottom of 115 ml serum bottles.After an overflow of at least 115 ml groundwater, the tubing was carefully removed and the bottles were immediately sealed with grey butyl rubber septa (ALT-MANN, Holzkirchen, Germany) and aluminium crimp caps.There were no visible air bubbles in the tubings and the vial during the procedure.The samples were stored at 10 • C (approximate groundwater temperature as estimated from mean annual air temperature) and analyzed within one week.Eight ml of Helium was injected in each vial in order to replace an equivalent amount of groundwater and to create a gas headspace.
Liquid and gas phase were equilibrated at constant temperature (25 • C) by agitating on a horizontal shaker for 3 h.To analyse N 2 and Ar, 1 ml headspace gas was injected manually with a gas-tight 1-ml syringe equipped with a valve (SGE, Darmstadt) into a gas chromatograph (Fractovap 400, CARLO ERBA, Milano) equipped with a thermal conductivity detector and a packed column (1.8 m length, 4 mm ID, molecular sieve 5 Å) and using helium as carrier gas.Because retention times of O 2 and Ar are similar on this column, O 2 was quantitatively removed using a heated Cu-column (800 flushed with helium immediately before penetrating the sample septum.Subsequently, the syringe was "over-filled" by approximately 15%, the syringe valve closed and the plunger adjusted to 1 mL in order to slightly pressurize the sample.The syringe needle was then held directly above the injection port before the valve was opened for a second to release excess pressure and the sample was finally injected.Generally, 3 replicate groundwater samples were analysed.A fourth sample served as reserve in case of failure during analysis.A calibration curve was obtained by injecting 0.2, 0.3, 0.5 and 1.0 ml of atmospheric air (3 replications each), resulting in different Ar and N 2 concentrations per calibration step.
To determine dissolved N 2 O concentrations, the headspace volume was augmented to 40 ml by an additional injection of 32 ml of Helium and an equivalent amount of groundwater was replaced.After equilibrating liquid and gas phase at constant temperature ( 25• C), 24 ml of the headspace gas were equally distributed to 2 evacuated septum-capped exetainers ® (12 ml, Labco, Wycombe, UK).Nitrous oxide was analyzed using a gas chromatograph equipped with an electron capture detector and an autosampler as described by Well et al. (2003).NO − 3 concentration was determined on 0.45 µm membrane-filtered samples by use of an ion chromatograph (ICS-90, DIONEX, Idstein, Germany) equipped with an IC-AIS column.
Molar fractions of N 2 , Ar and N 2 O in the headspace of sample vials and the volume of added He as well as the solubilities of these gases (Weiss, 1970(Weiss, , 1971;;Weiss and Price, 1980) were used to calculate partial pressure and molar fraction in the groundwater for each gas (Blicher-Mathiesen et al., 1998).Total pressure in the headspace after equilibration at 25 • C obtained from the sum of partial pressures of each gas or by direct measurement using a pressure transducer equipped with a hypodermic needle (Thies Klima, G öttingen, Germany) were in good agreement, i.e. differences between measured and calculated pressure were <9%.We checked the accuracy of estimated molar concentrations of dissolved gases from headspace concentration by adding defined volumes of N 2 (1 and 2 mL, respectively) to samples of demineralised water equilibrated at 10 • C. Recovery of N 2 was found to be satisfactory and was 92.91% for 1 Introduction

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Full Screen / Esc Printer-friendly Version Interactive Discussion and 2 mL added N 2 .

Calculation of excess N 2
N 2 dissolved in groundwater samples includes atmospheric N 2 and N 2 from denitrification (excess N 2 ) accumulated during the groundwater flow path (Boehlke, 2002).Principally, N 2 from denitrification can be determined by subtracting atmospheric N 2 from total N 2 (N 2T ).Atmospheric N 2 in groundwater consists of two components, (i) N 2 dissolved according to equilibrium solubility (N 2E Q ), and (ii) N 2 from "excess air" (N 2E A , Heaton and Vogel, 1981).Excess air denotes dissolved gas components in excess to equilibrium and other known subsurface gas sources.Excess air originates from entrapment of air bubbles at the groundwater surface during recharge which is subject to complete or partial dissolution (Holocher et al., 2002).
Excess N 2 (X excessN2 ) can thus be calculated using the following equation: where X denotes molar concentration of the parameters.X N2T represents the molar concentration of the total dissolved N 2 in the groundwater sample.X N2EQ is the molar concentration of dissolved N 2 in equilibrium with the atmospheric concentration.It depends on the water temperature during equilibration with the atmosphere, i.e. the temperature at the interface between the unsaturated zone and the groundwater surface.For the equilibrium temperature we assumed a constant value of 10 • C which was close to mean groundwater temperature.This is also similar to the mean annual temperature which is the best estimate of the mean temperature at the interface between unsaturated zone and the aquifer (Heaton and Vogel, 1981).X N2EQ was thus obtained using N 2 solubility data (Weiss, 1970) for this recharge temperature.N 2EA represents N 2 from excess air.For a given recharge temperature, excess air is reflected by noble gas concentrations (Holocher et al., 2002).If excess air results from complete dissolution of gas bubbles, the gas composition of the excess air component is identical to Introduction

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Full Screen / Esc Printer-friendly Version Interactive Discussion atmospheric air.For this case, X N2EA can be calculated from the concentration of only one noble gas, e.g.Argon (Heaton and Vogel, 1981): where X N2 atm and X Ar atm denote atmospheric mole fractions of N 2 and Ar, respectively.X Ar T represents the molar concentration of the total dissolved Ar in the groundwater sample.X Ar EQ is the molar concentration of dissolved Ar in equilibrium with the atmospheric concentration.
If excess air originates from incomplete dissolution of entrapped gas bubbles, then the N 2 -to-Ar ratio of excess air is lower than the atmospheric N 2 -to-Ar ratio due to fractionation (Holocher et al., 2002).The minimum value of the N 2 -to-Ar ratio of excess air is equal to the N 2 -to-Ar ratio in water at atmospheric equilibrium (Aeschbach-Hertig et al., 2002) since this value is approximated when the dissolution of entrapped air approaches zero.The minimum estimate of X N2 EA is thus given by where X N2 EQ and X Ar E Q denote equilibrium mole fractions of N 2 and Ar, respectively.
The actual fractionation of excess air can only be determined by analysing several noble gases (Aeschbach-Hertig et al., 2002).Because we measured only Ar, our estimate of excess N 2 includes an uncertainty from the unknown N 2 -to-Ar ratio of the excess air component.This uncertainty (U) is equal to the difference between N 2EA calculated with Eqs. ( 2) and (3), and is thus given by U N2 EA = (X Ar T − X Ar EQ ) × (X N2 atm /X Ar atm − X N2 EQ /X Ar EQ ) (4) It can be seen that U N2 EA directly depends on excess Ar, i.e.X Ar T −X Ar EQ .We used Eqs.
(1) to (3) to calculate minimum and maximum estimates of excess air and excess N 2 and assessed the remaining uncertainty of our excess N 2 estimates connected with excess air fractionation.Finally, we calculated means from the minimum and maximum values which we considered as best estimates of excess N 2 .Introduction

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Full Precision of the method was tested by evaluating standard deviation (σ) and repeatability (R).σ was determined for N 2 and Ar concentrations in atmospheric air samples (n=20), giving 0.000069 for Ar and 0.006449 for N 2 , respectively.Repeatability (R) was derived from R=2 √ 2 σ, giving 0.000196 for cAr (R Ar ) and 0.018241 for cN 2 (R N2 ).
Errors resulting from R N2 and R Ar were obtained using Eqs.(1-3), giving 1.59 and 2.05 mg N L −1 , respectively.Finally, total error for excess N 2 was determined by Gaussian error propagation giving 2.58 mg N L −1 for excess N 2 .
2.5 Initial NO − 3 concentration, reaction progress and emission factors NO − 3 input to a given spot of the aquifer surface is defined by the NO − 3 concentration of the seepage water or the groundwater directly at the groundwater table which is not yet altered by NO − 3 consumption by denitrification in the groundwater.In the following, this concentration is referred to as "initial NO − 3 concentration" (cNO − 3 t0 ).From the assumption that NO − 3 consumption on the groundwater flow path between the aquifer surface at a given sampling spot originates from denitrification and results in quantitative accumulation of gaseous denitrification products (N 2 O and N 2 ), it follows that cNO − 3 t0 can be calculated from the sum of residual substrate and accumulated products (B öhlke, 2002).Thus, cNO 3 -N t0 is given by the following equation: "Reaction progress" (RP) is the ratio between products and starting material of a process and can be used to characterize the extent of NO − 3 elimination by denitrification (B öhlke, 2002).RP is generally correlated with excess N 2 in denitrifying aquifers and is calculated as follows:

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Full "Emission factors" (EF) for indirect N 2 O emission from the aquifer resulting from Nleaching were calculated as described earlier (Well et al., 2005a).Because cNO represents the N-input to the aquifer via leaching, our data set is suitable to calculate an EF(1) from the relationship between N 2 O emission and N input, which is the ideal concept of emission factors (see introduction): Furthermore, we will compare EF(1) with the ratio of cN 2 O-N to cNO − 3 -N (EF(2)), which was used by the IPPC methodology (1997) to derive EF5-g.This concept was frequently used in recent studies to characterize indirect emissions in agricultural drainage water or groundwater (Reay et al., 2003;Sawamoto et al., 2005;) but it is non-ideal, because it assumes that these aquatic systems act solely as a domain of transport without any processing of NO − 3 and N 2 O (Well et al., 2005a, see introduction).The comparison between EF(1) and EF(2) will demonstrate potential errors in predicting indirect N 2 O emission from denitrifying aquifers using EF(2).

Basic groundwater properties, controlling factors O 2 and pH
Basic groundwater properties of the investigated aquifers are shown in Table 1.Groundwater temperatures were relatively constant at 10 • C. The pH and O 2 concentrations of the groundwater were more variable, suggesting heterogenous conditions for denitrification and N 2 O accumulation.The ranges of O 2 concentrations were similar in all aquifers and demonstrate that the investigated wells included both aerobic and anaerobic zones of each aquifer.Most of the sandy aquifers are acidic (Sulingen, Fuhrberg, Th ülsfelde) with similar pH ranges, whereas pH of the G öttingen gravel aquifer is close to 7.  2).Calculated initial NO − 3 concentrations (NO − 3 t0 , Eq. 5) were significantly higher than measured NO − 3 concentrations (Table 2), especially in the aquifer of Th ülsfelde.The difference between measured NO

N 2 O concentrations and emission factors
Wide ranges of N 2 O concentrations were observed in all aquifers (Fig. 1, Table 2).Highest concentrations up to 1271 µg N 2 O-N L −1 were measured in shallow groundwater at the Fuhrberg site at a RP of 0.3.
Emission factors EF(1) and EF(2) were highly variable (Table 3).Their medians for the complete data set were 0.00081 and 0.0031, respectively.Thus, EF(2) was in very good agreement with the 2006 IPCC default value for the EF5-g (IPCC, 2006), which was defined as 0.0025.In contrast, EF(1) was significantly lower than the 2006 IPCC default value.For each aquifer, EF(2) was substantially higher than EF(1).Within the sites, median values for each emission factor covered approximately one order of Introduction

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Full Screen / Esc Printer-friendly Version Interactive Discussion magnitude (EF(1): 0.00043 to 0.00438, EF(2): 0.00092 to 0.01801).For both EFs, we determined highest values for the Fuhrberg aquifer and lowest for the aquifer of G öttingen (Table 3).For the Fuhrberg and the Sulingen sites, we found EF(1) median values which are close to the 2006 IPCC default value of 0.0025.In contrast, we determined significant lower EFs(1) for the aquifers of Th ülsfelde and G öttingen.
N 2 O concentrations followed a rough pattern during RP.Values were lowest at the beginning (RP close to 0) and at the end (RP close to 1) but relatively high at a RP between 0.2 and 0.6 (Fig. 1).The same pattern was found for EF(1), which is strongly correlated to N 2 O concentrations (Table 4).However, at each RP we observed a relatively wide range of N 2 O concentrations and EF(1).

Uncertainty of excess N 2 estimates and excess N 2 related parameters
A certain amount of excess air, i.e. dissolved gas components in excess to equilibrium originating from entrapment of air bubbles at the groundwater surface during recharge (see Sect. 2.3), is often found in aquifers (Green et al., 2008).Although Heaton and Vogel (1981) assumed total dissolution of entrapped gas bubbles for their data set, fractionation of excess air (that means partial solution of the bubbles) is a probable phenomenon (see Sect. 2.3).This was clearly shown by Aeschbach-Hertig et al. (2002) for different aquifers and different environmental conditions.The extent of fractionation of excess air could not be assessed in our data set, because this requires analysing of several noble gases, what was not done in this study.Therefore, we used the means of minimum and maximum values for excess N 2 as a possible estimate which were calculated assuming complete dissolution or maximum fractionation of entrapped gases, respectively (see Sect. 2.3, Eqs. 2 and 3).The maximum error is thus half the difference between minimum and maximum estimates.The uncertainty connected with this procedure is documented in Fig. 2, where "excess N 2 min" and "excess N 2 max" Introduction

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Full Screen / Esc Printer-friendly Version Interactive Discussion denote minimum and maximum estimates for excess N 2 , respectively.Derived from the whole data set shown in Fig. 2, the mean difference between minimum and maximum estimates for excess N 2 is 1.25 mg N L −1 and the mean of the maximum errors is thus 0.63 mg N L −1 .According to Eq. ( 5), these error values are also valid for NO − 3 t0 .Using the uncertainty of excess N 2 and NO − 3 t0 we also estimated the uncertainty of RP (Eq.6), giving 0.008 for the mean of the maximum errors.This shows that the uncertainty of RP has only little implication of our conclusion that maximum N 2 O concentrations occured at RP between 0.2 and 0.6 and for the relationship between RP and emission factors shown in Fig. 3. From Eq. ( 7) it follows that the relative error of EF( 1) is equal to the relative error in NO − 3 t0 , giving 4.8% for the median NO − 3 t0 of 13.15 mg N L −1 .In view of the large range of EF(1) (Table 3) this uncertainty is small.
Therefore, it can be concluded that the consequences of uncertainties connected with excess N 2 and NO − 3 t 0 are negligible for our concept of EF(1).Significant degassing of groundwater may occur when the sum of partial pressures of dissolved gases (e.g.Ar, N 2 , O 2 , CO 2 , and CH 4 ) exceeds that of the hydrostatic pressure.This phenomenon was found when high denitrifying activity induced production of excess N 2 in shallow groundwater of riparian ecosystems (Blicher-Mathiesen et al.,1998;Mookherji et al., 2003).In our study, the sum of partial pressures never exceeded hydrostatic pressure which is in part due to the fact, that the majority of data originates from deeper groundwater (Table 1) where hydrostatic pressure is higher than in upper groundwater.These conditions prevent degassing of gaseous denitrification products.Water samples from shallow groundwater, where the risk of degassing is higher due to lower hydrostatic pressure, were only taken from the Fuhrberg site.denitrification progress during aquifer passage.These relationships and additional significant positive correlations between sampling depth and excess N 2 were mostly pronounced in the partial data-set of Fuhrberg, whereas the correlations were lower or insignificant for the other aquifers (data not shown).The latter suggests that spatial distribution of denitrification within these aquifers was more heterogeneous which implies that the relationship between reaction progress and residence time was more variable.A significant negative correlation between NO − 3 and excess N 2 in the whole data-set (R S =−0.37,Table 4) demonstrates that denitrification was an important factor for NO − 3 variability within all aquifers.With increasing NO − 3 concentration the N 2 O-to-N 2 ratio may strongly increase (Kroeze et al., 1989) because NO − 3 usually inhibits N 2 O reduction to N 2 (Blackmer and Bremner, 1978;Cho and Mills, 1979).This is confirmed by the positive correlation between N 2 O and NO − 3 we evaluated in this study (Table 4).A significant negative correlation was found between N 2 O and pH, which was mostly pronounced in the aquifer with the widest pH range (Fuhrberg, see Table 1, spearman correlation coefficient (R S )=−0.33).N 2 O accumulation in aquifers might be supported by increasing groundwater acidity because the reduction step of N 2 O to N 2 is much more sensitive to acidic conditions compared to the preceding reduction steps (Granli and Bøckman, 1994).This regulation is illustrated by the negative correlation between pH and N 2 O in our study.The influence of pH on the N 2 O/N 2 ratio is intensified by high NO − 3 concentrations (Blackmer and Bremner, 1978;Firestone et al., 1980).Due to these observations we conclude that conditions were especially favourable for N 2 O accumulation and potential N 2 O emission in shallow groundwater of the Fuhrberg aquifer, because it is characterized by high NO − 3 contamination and comparatively low pH.This is confirmed by our data since N 2 O concentrations of these samples were highest within the entire data-set.

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Full decrease during further residence time in the aquifer or during the passage of the unsaturated zone before it reaches the atmosphere.Moreover, diffusive N 2 O emission from the aquifer surface to the unsaturated zone and eventually to the atmosphere (Deurer et al., 2007) is not taken into account by EF(1).Therefore, the measured data supply only potential emission factors quantifying the amount of N 2 O which could be emitted, if the groundwater was immediately discharged to springs, wells or streams.The determination of an effective emission factor to quantify real N 2 O flux from the investigated aquifers requires validated models of reactive N 2 O transport.Further research on reaction dynamics and gas transport within the aquifers is needed to achieve this.
However, the comparison of N 2 O concentration and EF(1) with RP gives a rough sketch of the principal N 2 O pattern during groundwater transport through denitrifying aquifers.Although variations of N 2 O and EF(1) at any given level of RP was high, there was a clear tendency of low N 2 O concentrations for RP close to zero or close to 1 and highest N 2 O concentrations at RP between 0.2 and 0.6.This pattern is consistent with the time course of N 2 O during complete denitrification in closed systems observed by modelling (Almeida et al., 1997) as well as laboratory incubations (Well et al., 2005b) and can be explained by the balance between production and reduction of N 2 O during a Michaelis-Menten reaction kinetics.It can be concluded that RP can be considered as an important parameter to predict N 2 O emission via groundwater discharge.This emission can be expected to be negligible if RP at groundwater discharge is very small or close to 1. Conversely, relatively high emission can be expected if RP at groundwater discharge is between 0.2 and 0.6.The observed relationship suggests, that emission Introduction

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Full factors are also related to denitrification rate, groundwater residence time and sampling depth because these quantities determine the reaction progress.This could be helpful to predict or interpret N 2 O emission from different types of groundwater systems.For example, low N 2 O fluxes observed from tile drainage outlets (Reay et al., 2003) might be explained by relatively low groundwater residence time of this drainage system.The deep wells of the investigated aquifers with low residual NO − 3 and low N 2 O concentration reflect the typical low emission factors at RP close to 1. Hot spots of N 2 O emission from groundwater might be locations were groundwater is discharged to surface waters immediately after partial NO − 3 consumption which is known to occur after the subsurface flow through riparian buffers (Hefting et al., 2003).
A downward revision of the EF5-g default value by the IPCC from 0.015 (1997) to 0.0025 ( 2006) was based on recent findings of Hiscock et al. (2002Hiscock et al. ( , 2003)), Sawamoto et al. (2005) and Reay et al. (2005).This is supported by site medians of EF(1) of this study (Table 3) which scatter around the revised EF5-g.Obviously, the former 1997 IPCC EF5-g default value of 0.015 substantially overestimated indirect N 2 O emissions from groundwater.A comparison of the emission factors EF(1) and EF(2) clearly shows lower values for EF(1) which results from the consideration of initial NO − 3 by EF(1).The deviation between EF(1) and EF(2) is highly relevant in aquifers with substantial denitrifying activity and high N inputs like those investigated in this study.Furthermore, Fig. 3 demonstrates that differences between EF(1) and EF(2) are increasing with reaction progress of denitrification.This clearly demonstrates that it is important to take the dynamic turnover of NO − 3 during groundwater passage into account.Consequently, potential N 2 O emissions from aquifers should be estimated using EF(1) rather than EF(2).
). H öll et al. (2005) stated that these transformations are the reason why N 2 O concentration in groundwater does not necessarily reflect actual indirect N 2 O emission.Finally, as a result of NO − 3 consumption in denitrifying aquifers, the NO − 3 concentration in the deeper groundwater is lower than the initial NO − 3 concentration at the groundwater surface.Thus, the reconstruction of initial NO deviation and repeatability of excess N 2 analysis an important factor in all investigated aquifers.
Unlike the observations of Blicher-Mathiesen et al. (1998) and Mookherji et al. (2003) excess N 2 in the shallow groundwater measured in this study was relatively low and hydrostatic pressure was thus not exceeded by accumulation of dissolved gases.The fact that calculation of initial NO − 3 concentration is based on excess N 2 implies a need for quantitative estimates of excess N 2 in order to determine EF(1) accurately.But it also involves the possibility to validate excess N 2 in cases where NO − 3 t0 is known.
determined from measured fluxes across the soil surface, emission factors estimated from groundwater concentration do not reflect the actual N 2 O emission from the system because the amount of dissolved N 2 O might increase or the groundwater surface by adding up concentrations of NO − 3 , N 2 O and excess N 2 .Because this initial NO − 3 concentration reflects the N input to the groundwater by leaching it was used to calculate an emission factor EF(1) for indirect agricultural N 2 O emissions from groundwater which is for the first time based on the ratio between N 2 O concentration and N-input.An uncertainty of excess N 2 estimates according to the excess air phenomenon was found to be negligible for this concept of EF(1).EFs(1) in the investigated denitrifying aquifers were much lower than the values resulting from the earlier concept of groundwater emission factors consisting of N 2 O-to-NO − 3 ratios of groundwater samples (EF(2) in this study).This demonstrates the need to take past NO − 3 consumption into account when determining groundwater emission factors.In agreement with recent literature data our observations support the substantial downward revision of the IPCC default EF5-g from 0.015 (1997) to 0.0025(2006).However, there are still uncertainties with respect to a single emission factor for the effective N 2 O flux from the investigated aquifers because spatial und temporal heterogeneity of N 2 O concentrations was high and further metabolism of N 2 O during transport in the aquifer and through the unsaturated zone before it is emitted is poorly understood

FuhrbergFig. 1 .
Fig. 1.N 2 O in groundwater samples from 4 different aquifers in relation to reaction progress.Reaction progress is the ratio between denitrification products (excess N 2 +N 2 O) and initial NO − 3 .
Ranges and site medians of reaction progress and excess N 2 are given in Table2.Lowest values for excess N 2 coincided with RP of approximately 0. A RP of approximately 1 was characterized by high values of excess N 2 in all aquifers.In all aquifers, samples cover almost the complete range of RP.Highest excess N 2 values were observed at Th ülsfelde, which were twice the values of the other sites.At the drinking water well of the Fuhrberg catchment, NO

Table 1 .
General properties for the aquifers of Fuhrberg, Wehnsen, Sulingen, Th ülsfelde and G öttingen.Weymann et al.: "Assessment of Excess N 2 and Groundwater N 2 O Emission Factors of Nitrate-contaminated Aquifers in Northern Germany", Figure 1.