(N2O + N2) product ratio of denitrification as controlled by available carbon substrates and nitrate concentrations

Agriculture, Ecosystems and Environment 147 (2012) 4–12

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N2 O emission and the N2 O/(N2 O + N2 ) product ratio of denitrification as controlled by available carbon substrates and nitrate concentrations M. Senbayram a,b,∗ , R. Chen c , A. Budai d , L. Bakken d , K. Dittert b,e a

Institute of Plant Nutrition and Environmental Science, Research Center Hanninghof, Yara Int. ASA, Duelmen, Germany Institute of Plant Nutrition and Soil Science, Christian-Albrechts University, Hermann Rodewald Str.2, Kiel, Germany State Key Laboratory of Soil and Sustainable Agriculture, Institute of Soil Science, Chinese Academy of Sciences, 210008 Nanjing, China d Institute of Plant and Environmental Sciences, Norwegian University of Life Sciences, As, Norway e Department of Crop Sciences, Plant Nutrition and Crop Physiology, Georg-August-University, Goettingen, Germany b c

a r t i c l e

i n f o

Article history: Received 9 October 2010 Received in revised form 17 June 2011 Accepted 23 June 2011 Available online 22 July 2011 Keywords: Labile carbon Nitrate Nitrous oxide emission Dinitrogen emission Denitrification

a b s t r a c t Amending agricultural soils with organic residues is frequently recommended to improve soil fertility and to sequester carbon for counteracting global warming. However, such amendments will enhance microbial respiration, hence denitrification. Therefore, the assessment of effects on global warming must take N2 O emission and the N2 O/(N2 O + N2 ) product ratio of denitrification into account. There are some indications that the product ratio of denitrification is positively correlated with the ratio of available NO3 − and available organic C in soils, but more research is needed to unravel quantitative relationships in well defined experiments. We conducted two laboratory incubation experiments, with the objective (i) to test the impact of the application of various N containing organic substrates including biogas residue on the denitrification rate and on N2 O emission, and (ii) to investigate the effect of various NO3 − concentrations on the denitrification rate and the N2 O/(N2 O + N2 ) product ratio under standardized anoxic conditions in soils collected from long-term organic or inorganic fertilizer plots. In experiment 1, we found that biogas residue was more recalcitrant than maize straw, despite a high concentration of soluble organic C. High respiration (treatments with maize straw and sucrose) resulted in a transient peak in N2 O emission, declining rapidly towards zero as nitrate concentrations reached less than 20 mg NO3 − -N kg−1 dry soil. Application of biogas residue had a more moderate effect on soil respiration and denitrification, and resulted in a more long lasting peak in N2 O emission. The results were interpreted as a result of a gradual increase in the relative activity of N2 O reductase (thus lowering of the N2 O/(N2 O + N2 ) product ratio of denitrification) throughout the incubation, most likely controlled by concentration of available NO3 − in soil. In the second experiment, we found low N2 O/(N2 O + N2 ) product ratios for the treatment where NO3 − concentrations were ≤2 mM, and the ratios were clearly lower in manure fertilized than in mineral fertilizer treated soil. Much higher N2 O/(N2 O + N2 ) product ratios were found for the treatments with ≥10 mM NO3 − , and the ratios were remarkably independent of the soil’s fertilizer history. We conclude that (i) in N-fertilized agricultural soils, application of organic matter with high contents of labile C may trigger denitrification-derived N2 O emission whereas (ii) in soils with low NO3 − contents such application may substantially lower the N2 O/(N2 O + N2 ) product ratio and hence N2 O emission. © 2011 Elsevier B.V. All rights reserved.

1. Introduction As a potent greenhouse gas, nitrous oxide (N2 O) is responsible for about 6% of the current greenhouse effect (Bouwman et al., 1995; IPCC, 2007). Moreover, N2 O has recently received great attention because of its importance for stratospheric ozone depletion (Ravishankara et al., 2009).

∗ Corresponding author Present address: Institute of Plant Nutrition and Environmental Science, Research Center Hanninghof, Yara Int. ASA, Duelmen, Germany. E-mail address: [email protected] (M. Senbayram). 0167-8809/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.agee.2011.06.022

Globally, agricultural soils account for about 60% of the atmospheric N2 O emissions (Mosier et al., 1998; Kroze et al., 1999). Biological processes, especially nitrification and denitrification are the major sources of N2 O emissions from agricultural soils. Denitrification is a microbially mediated process of dissimilatory nitrate reduction that may produce nitric oxide (NO), N2 O and molecular nitrogen (N2 ). The reduction of nitrate to nitrite (E0 = +420 mV), of nitrite to nitric oxide (E0 = +375 mV), of nitric oxide to nitrous oxide (E0 = +1175 mV) and finally to nitrogen gas (E0 = +1355 mV) are catalyzed by specific enzymes (Zannoni, 2004). In situ, denitrification rates depend on oxygen availability, soil moisture, soil type, pH, nitrate concentration, and quite impor-

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Table 1 Summary of soil characteristics in Hohenschulen soil, in long-term mineral N treated sandy loam soil (LMN), long-term organic manure treated sandy loam soil (LOM), and in Lavesum soil (SL).

Hohenschulen (Exp. 1) LMN (Exp. 2) LOM (Exp. 2) SL (Exp. 2)

Total N %

Total C %

pH

WHC w/w

Sand %

Silt %

0.11 0.05 0.07 0.16

1.5 0.8 1.0 2.0

6.5 6.4 6.4 5.5

31 33 35 38

58 83 83 34

29 10 10 58

tantly on the availability of labile carbon compounds in soil (Burford and Bremner, 1975; Dittert et al., 2005; Loecke and Robertson, 2009; Senbayram et al., 2009). Plant residues and animal manures represent the main inputs of carbon compounds to agricultural soils. Recently, bioenergy by-products, e.g. biogas residues, are increasingly used as energy production by biomass fermentation grows rapidly. Long-term soil amendment with organic substrates may efficiently increase soil carbon content (Stratfon and Rechcigl, 1998; Meng et al., 2005) contributing to soil fertility and carbon sequestration in soil. Cayuela et al. (2010) reported that for this purpose, organic fertilizers should contain stable C compounds, as otherwise they would be vulnerable to rapid microbial decomposition in soil. Furthermore, soil amendment with organic fertilizers that contain readily-decomposable organic carbon compounds may trigger denitrification by enhancing respiration (through creation of anoxic micro-sites) and by providing energy for denitrifiers (Burford and Bremner, 1975; Firestone, 1982; Weier et al., 1993). Therefore, environmental benefits from amending soils with organic fertilizers may be counterbalanced depending on the extent of N2 O emission being induced by enhanced denitrification. As nitrate (NO3 − ) is very mobile in soil, it may rapidly diffuse into soil compartments with low oxygen contents where it may promote biological denitrification. Thus, next to degradable carbon compounds, the NO3 − concentration in the soil solution is another major factor limiting denitrification. For example, Luo et al. (1998) observed that the rate of diffusion of NO3 − to denitrification micro-sites limited denitrification in grazed pasture. Therefore, the application of nitrogen fertilizers or manures increases denitrification rates especially when there is an ample supply of carbon (Lampe et al., 2006; Fangueiro et al., 2008; Bhandral et al., 2010). However, higher denitrification rates do not necessarily lead to higher N2 O losses, since enhanced denitrification may also alter the share of N2 O among the two gaseous products N2 O and N2 , i.e. the N2 O/(N2 O + N2 ) product ratio. It has repeatedly been observed that this ratio is affected not only by the availability of C but also by the ratio of nitrate and available C in arable soils (Weier et al., 1993; Van Cleemput, 1998), in tropical forests (Kiese and ButterbachBall, 2002), or in laboratory incubation experiments (Alinsafi et al., 2008). However, direct measurements of molecular N2 gas have not been done in these studies, and its flux was indirectly estimated, e.g. deduced from reduction in nitrate concentrations, or by inhibition of N2 O reductase through acetylene. In addition, the combined effects of the relevant variables on the N2 O/(N2 O + N2 ) product ratio are still not well understood. We set up two incubation experiments under fully controlled conditions. The objective of the first experiment was to test the impact of the application of various N-containing organic substrates on the denitrification rate and on N2 O emission. We compared biogas residue (as low available C source) with maize straw (as C source with moderate availability), and sucrose (highly available C source), aiming at understanding effects of the quantity and quality of added carbon sources on N2 O emission from denitrification. In the second experiment, we investigated the effects of various NO3 − concentrations on the denitrification rate and the N2 O product ratio of denitrification in soils with long-term application of organic or inorganic fertilizer.

Clay % 13 7 7 8

2. Materials and methods 2.1. Pot experiment (Exp. 1) 2.1.1. Experimental design The soil used in the pot experiment was collected in spring 2008 at Hohenschulen experimental farm of Kiel University, (10.0◦ E, 54.3◦ N) 15 km west of Kiel, northern Germany. Prior to the experiment, arable crops (oilseed rape, wheat, barley) were grown on this soil, which is classified as Stagnic Luvisol with a sandy loam texture (Table 1). For water holding capacity (WHC) measurement, the soil samples were saturated with water in a PVC cylinder (30 cm height). The cylinder was placed on an absorbent membrane until all excess water was drawn away by gravity. Once equilibrium was reached, the water holding capacity was calculated based on the weight of the water held in the sample vs. the sample dry weight. A more detailed description of the soil is given by Horn (2005). Before use, the soil was air-dried (to 30% of WHC ± 2%), homogenized and passed through a 2 mm sieve. Glass pots with closed bottom, 250 ml in volume and 12 cm in height were used, and each pot was equipped with an air-tight lid. The aim of the experiment was to simulate good agricultural practices for organic substrate application, i.e. immediate incorporation of organic substrate into the top soil layer. Prior to the experiment, soils were pre-incubated for 10 days at 45% WHC to allow microbial activity to become stabilized. Soil moisture was adjusted to 75% WHC, equivalent to 67% water filled pore space (WFPS) in order to create semi-anaerobic conditions (Dobbie and Smith, 2001). The experimental treatments consisted of four types of organic amendments, all added together with ample amounts of NH4 -N (110 mg NH4 + -N kg−1 dry soil, equivalent to 150 kg NH4 + N ha−1 ), and two control treatments (with and without NH4 + , see Table 2 for details). The purpose of adding similar amounts of ammonium to all treatments was to secure sufficient ammonium concentrations throughout the entire incubation in all treatments hence avoiding microbial assimilation of NO3 − . The soil in each pot (150 g dry weight) was mixed with a suspension of the respective fertilizer and water and immediately filled into pots. The same physical mixing of the soil was done with the unamended control soil. Considering water-content in the organic manure, water was added to reach equal levels in the final soil moisture of 75% of WHC. Biogas residue, maize straw, sucrose and maize straw + sucrose were all supplied as slurry solution, and the mineral fertilizer (ammonium sulfate) was also added as a solution. Experimental treatments and the amount of N and C applied are given in Table 2. In the maize straw treatment (MST) the added amount of total organic carbon (2.5 g C kg−1 soil) was chosen to be similar to the biogas residue treatment (BGR) whereas in sucrose treatment (SUC),

Table 2 Total C (g C kg−1 dry soil) and N (mg NH4 + -N kg−1 dry soil) additions in treatments of control, mineral-N (MIN), biogas residue (BGR), maize straw (MST), sucrose (SUC), and maize straw + sucrose (MST + SUC) in Exp. 1.

Total carbon NH4 + -N

Control

MIN

BGR

MST

SUC

MST + SUC

0 0

0 110

2.5 110

2.5 110

0.5 110

3.0 110

6

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sugar was applied at only 1/5 of the total C dose (0.5 g C kg−1 soil) of straw and biogas residue. Biogas residue originated from an agricultural biogas plant in Marienthal, Germany, where ensiled maize plants were fermented at 40 ◦ C for 65 days. The fermentation effluent collected from the storage tank was used directly without further treatment. The key properties of the biogas residue were 31% total C, 5.5% total N and a pH of 7.8. In MST and BGR, the added amount of organic N were 79 and 128 mg NH4 + -N kg−1 dry soil respectively. Each treatment was done in six replications, from which three were used for gas sampling and the other three for soil sampling. After fertilizer application, soils were incubated for 124 h, keeping an air temperature of 19 ◦ C and relative air humidity of 65%. 124 h after on-set of the treatments, final water contents were 65% of WHC. 2.1.2. Soil NH4 + , NO3 − and dissolved organic carbon (DOC) concentrations For the analysis of soil mineral N, soils were sampled and extracted with 2 M KCl solution (1:4 w/v) by shaking for 1 h. The extracts were then filtered through Whatman 602 filter paper and stored at −20 ◦ C until analysis. The concentrations of NH4 + and NO3 − in soil extracts were measured colorimetrically using a TRAACS 800 autoanalyzer (Bran and Luebbe, Norderstedt, Germany). Dissolved organic carbon (DOC) and dissolved inorganic carbon (DIC) were analyzed according to Cookson et al. (2007). Briefly, soil samples were extracted with a 0.5 M K2 SO4 solution (1:4 w/v) by shaking for 1 h. The extracts were then filtered through Whatman 602 filter paper and immediately measured using a DIMATEC TC/TIC/TOC analyzer (Dimatec, Essen, Germany). 2.1.3. Trace gas measurement techniques Using a closed chamber method (Hutchinson and Mosier, 1981), gas emission was measured every 6 h from each pot during the first 36 h of incubation. Thereafter, emissions were measured every 12 or 24 h over the following 54-h period. Before each emission measurement, the room was ventilated for 15 min to homogenize the ambient air background. Then, PVC lids were fitted onto each pot. 0, 20 and 40 min after sealing, 10 ml of gas were sampled using pre-evacuated Exetainer glass bottles (Labco, High Wycombe, UK). N2 O and CO2 concentrations were analyzed by ECD gas chromatography (Model 3400 CX, Varian Inc., Palo Alto, CA, USA). Operating conditions for the GC were as follows: injector temperature 95 ◦ C, column temperature 85 ◦ C, and detector temperature 320 ◦ C. Samples were introduced using a Gilson 222 XL autosampler (Gilson Inc., Middleton, USA). Data processing was performed using the Varian Star Chromatography software (ver. 6.2). 2.2. Robotized incubation experiment (Exp. 2) 2.2.1. Experimental design For this experiment, soils were collected in spring 2010 at Hanninghof long-term experiment Duelmen (north-western Germany, 51◦ 50 N; 7◦ 15 E). This long-term trial was established in 1958 with potato and winter-rye cultivation. More detailed information can be found in Melkamu (2010). Initially in 1958, the soil contained 1% total C, 0.1% total N and had a pH of 6 with a sandy-loam texture. Since 1958 both, mineral-N (LMN) and organic-N treatments (LOM) received mineral fertilizer as calcium–ammonium–nitrate (170 kg N ha−1 y−1 ). However, in the organic fertilizer plots, cattle or pig slurry was applied every third year at an average rate of 141 kg N ha−1 instead of mineral fertilizer. Soil parameters analyzed in 2010 are presented in Table 1. The third soil (SL) was collected on a farm in Lavesum (20 km west of Duelmen). This soil contained

slightly more clay and substantially more silt than LMN and LOM. A more detailed description of the Lavesum soil is given by Blankenau et al. (2000). Prior to the experiment, arable crops (maize, wheat, potato and barley) had been grown on this soil for several years. Before use, all soils were air-dried (to 30% WHC ±2%), homogenized, passed through a 2 mm sieve, and then wetted again to 65% WHC prior to the experiment for 10 days. The incubation experiment was done to determine effects of NO3 − concentration on potential denitrification and its N2 O/(N2 O + N2 ) product ratio when soil was incubated under complete anoxic conditions. Pre-wetted soils (LOM, LMN and SL) were treated with 0.2, 2, 10, and 20 mM KNO3 solution prior to the experiment. Briefly, 60 g moist soil was placed in PVC pots with porous ceramic plates at the bottom. They were then flooded with nitrate solution and drained to ca. 65% WHC using vacuum. Flooding and draining were repeated twice in order to reach a homogenous nitrate concentration and well-drained conditions for all soils. Then, the soil was immediately transferred to 120-ml serum flasks. Flasks were sealed with air-tight, butyl-rubber septa and aluminium crimp caps and made anoxic by repeated evacuation and filling with He. Triplicate samples were prepared for each treatment. In addition to block N2 O reduction to N2 , another set of triplicate samples was treated with acetylene (C2 H2 ) which was added to the headspace (through injection and subsequent release of over pressure, reaching a final concentration of 10%, v/v). C2 H2 treatments were used to crosscheck the reliability of the N2 measurements. 2.2.2. Gas flux measurements N2 O, NO and N2 production during batch incubation of soils was monitored by frequent (every 2 or 3 h) headspace analysis of the incubation bottles. We used an automated incubation system consisting of a thermostated water bath (15 ◦ C) with positions for 42 serum bottles (120 ml) with an autosampler (CTC PAL). Headspace gas was sampled to the GC via a Gilson Minipuls 3 peristaltic pump connected to a loop injection system by 1/16 steel tubing except for a small piece of thick-walled Marprene tubing in the pump (Watson Marlow, England). The principle of the robotized incubation system is described in detail by Molstad et al. (2007). The current experiment was conducted with a modified version of the system where the micro-GC had been replaced with a regular GC (Agilent). Headspace gas was sampled and analyzed automatically by an autosampler (CTC PAL) and a Gilson Minipuls3 peristaltic pump attached to a gas chromatography system (Agilent, Italy). The gas chromatograph (GC) was equipped with two columns; a 10 m PorapLOT U for N2 O and CO2 analysis (oven temperature 36 ◦ C) and a 20 m 5A Molsieve (oven temperature 50 ◦ C) for N2 and O2 analysis. The GC was equipped with an electron capture detector (for measuring low concentrations of N2 O), thermal conductivity detector (for measuring CO2 and high N2 O concentrations) and a flame ionization detector (for measuring CH4 ). The outlet of the GC injection loop was coupled to a T-piece junction with a continuous He flow, which carried the sample gas into a chemoluminescence NOx -analyzer (Model 200A, Advanced Pollution Instrumentation, San Diego, CA) for determination of NO concentrations. After each gas sampling, the peristaltic pump was reversed, thus replacing the sampled gas with an equal volume of He. Thus, the sampling did not alter the headspace gas pressure but led to a significant dilution (1%) of the headspace gas which was taken into account when calculating rates of gas production. 2.3. Data processing and statistical analysis Cumulative N2 O and CO2 emissions (Exp. 1) were calculated by linear interpolation between measured daily fluxes. Emission rates were expressed as arithmetic means of the three replicates and log-

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transformed for statistical analysis. In Exp. 2, kinetics of NO, N2 O and N2 accumulation during the incubation period were used to calculate the N2 O product share of denitrification and total denitrification rate. Mean denitrification rates as well as N2 O product shares of denitrification were calculated for the initial phase of the experiment when concentrations of N2 and N2 O increased linearly. The linear phase was shorter at low N rates than at high N rates due to the rapid depletion of NO3 − (calculated by mass balance, a large fraction of the NO3 − is recovered as N2 ). Tukey’s HSD posthoc test was used to reveal significant pairwise differences among treatments. Statistical analyses were done using SPSS version 13.0 (SPSS Inc., Chicago, IL, USA), with p < 0.05 used as the criterion for statistical significance. 3. Results and discussion 3.1. Experiment 1 3.1.1. CO2 emissions All added organic substrates induced a significant increase in respiration, but the time course of this stimulation varied greatly among substrates (Fig. 1A). Biogas residue immediately enhanced respiration which then gradually declined. In contrast, the other substrates induced a somewhat slower increase in respiration with peak rates after 40–70 h. The immediate response in BGR may be attributed to the activity of a significant bacterial population in the biogas residue itself, established during storage after bio-methane production. In contrast, the large inputs of easily degradable organic residues (sucrose and maize) induced increasing respiration, reflecting substantial microbial growth on this substrate. This is also reflected in the significant decrease of the soil NH4 + pool in these treatments (Fig. 2A). Cumulative CO2 emission in BGR was significantly lower than in MST (Table 3). Assuming that the mineralization of soil organic carbon is unaffected by the amendments (i.e. no priming effect), cumulative CO2 losses for the entire incubation can be used to calculate the approximate fraction of the added carbon substrates which are mineralized during the incubation. The mineralization of the substrate-C can then be estimated as the difference between cumulative CO2 -C evolved in amended soil minus that in the control soil. The calculated share of mineralized biogas residue-C and maize straw-C was 11 and 22% respectively. The low percentage of carbon mineralization in biogas residues is attributable to lower contents of available C compounds in this material, because the biogas fermentation process consumes most easily hydrolysable fractions in the organic substrates, and only the most recalcitrant compounds remain. 3.1.2. Soil NH4 + , NO3 − and dissolved organic carbon (DOC) concentrations Application of organic and inorganic fertilizers caused immediate increases in NH4 + concentrations, reaching a mean maximum of 140 mg NH4 + -N kg−1 dry soil in BGR (Fig. 2A). The concentration of NH4 + remained relatively constant throughout the incubation in the BGR and MIN treatments, suggesting a moderate nitrification rate possibly balanced by an equal rate of net N mineralization.

Fig. 1. Time course fluxes of (A) CO2 (mg CO2 –C kg−1 dry soil h−1 ), and (B) N2 O (mg N2 O–N kg−1 dry soil h−1 ) from soil after application of mineral-N (Min; at an N rate of 150 kg NH4 + -N ha−1 as ammonium-sulfate), biogas residue (BGR; at N rate of 150 kg NH4 + -N ha−1 ), maize straw + mineral-N (MST; at an N rate of 150 kg NH4 + N ha−1 ), sucrose + mineral-N (SUC; at an N rate of 150 kg NH4 + -N ha−1 ), and maize straw + sucrose + minerial-N (MST + SUC; at an N rate of 150 kg NH4 + -N ha−1 ) during 90 h of the pot experiment (Exp. 1). Error bars show the standard error of the mean of each treatment (n = 3). In some cases error bars are smaller than the symbols.

In contrast, soil amendment with easily decomposable substrates (maize straw and sucrose) resulted in substantial reduction of the ammonium concentration throughout the incubation. This contrast is attributable to net microbial assimilation of ammonium in response to the high C/N ratio of the assimilated substrates (maize straw and sucrose). When mineral N serves as an N source for microbial growth, it is preferentially assimilated as NH4 + rather than NO3 − , since the assimilation of the latter is strongly repressed by available NH4 + (Brown, 1980; Vallino et al., 1996; Geisseler et al., 2009). The fact that none of the treatments resulted in a depletion of NH4 + suggests that the observed rates of NO3 − depletion were due to denitrification rather than microbial assimilation of NO3 − . The soil NO3 − background concentration was 42 (±1.2) mg NO3 − -N kg−1 dry soil for all treatments prior to the experiment including the non-fertilized control soil. In all treatments, NO3 − concentrations decreased significantly throughout the incubation period (Fig. 2B). Final NO3 − concentrations were 0, 3, 12, 16 and 28 mg NO3 − -N kg−1 dry soil in MST + SUC, MST, SUC, BGR, and MIN respectively. Assuming that if at all there was only insignif-

Table 3 Cumulative emissions of CO2 (mg CO2 -C kg−1 dry soil), and N2 O (mg N2 O-N kg−1 dry soil) from soil after application of mineral-N (MIN; at an N rate of 150 kg NH4 + -N ha−1 as ammonium-sulfate), biogas residue (BGR; at an N rate of 150 kg NH4 + -N ha−1 ), maize straw + mineral-N (MST; at an N rate of 150 kg NH4 + -N ha−1 ), sucrose + mineral-N (SUC; at an N rate of 150 kg NH4 + -N ha−1 ), and maize straw + sucrose + mineral-N (MST + SUC; at an N rate of 150 kg NH4 + -N ha−1 ) during 90 h of the pot experiment (Exp. 1). Standard errors are given in brackets. Values with different letters are statistically different at p < 0.05.

N2 O CO2

Control

MIN

BGR

MST

SUC

MST + SUC

14.1 (±1.5)c 54.2 (±9)D

14.3 (±2.4)bc 56.9 (±6)D

16.3 (±0.5)bc 333.1 (±31)C

18.9 (±0.8)ab 625.4 (±9)B

22.5 (±3.9)a 588.3 (±17)B

17.9 (±1.1)bc 869.4 (±30)A

8

M. Senbayram et al. / Agriculture, Ecosystems and Environment 147 (2012) 4–12

Fig. 3. Cumulative CO2 emission versus net NO3 losses 124 h after the start of the experiment (Exp. 1). Error bars show the standard error of the mean of each treatment (n = 3). In some cases error bars are smaller than the symbols.

Fig. 2. Kinetic of (A) NH4 + (mg NH4 + -N kg−1 dry soil), (B) NO3 − (mg NO3 -N kg−1 dry soil), and (C) dissolved organic carbon (mg DOC–C kg−1 dry soil) concerntration from soil after application of mineral-N (Min; at an N rate of 150 kg NH4 + -N ha−1 as ammonium-sulfate), boigas residue (BGR; at an N rate of 150 Kg NH4 + -N ha−1 ) maize straw + mineral-N (MST; at an N rate of 150 kg NH4 + -N ha−1 ), sucrose + mineral-N (SUC; at an N rate of 150 kg NH4 + -N ha−1 ), and maize straw + sucrose + mineral-N (MST + SUC; at N rate of 150 kg NH4 + -N ha−1 ), during 124 h of the pot experiment (Exp. 1). Error bars show the standard error of the mean of each treatment (n = 3). In some cases error bars are smaller than the symbols.

rate, since higher denitrification rates were measured in MST than in BGR while both treatments received similar amounts of total C. In addition, these results reinforce the idea that soil amendment with organic fertilizers that contain readily decomposable organic carbon compounds may trigger denitrification by enhancing respiration (through consumption of oxygen creating anoxic micro-sites) and by providing energy for denitrifiers (Firestone, 1982; Weier et al., 1993; Attard et al., 2011). The concentrations of dissolved organic carbon (DOC) were about 51 (±2.5) mg C kg−1 dry soil, and they decreased slightly over time in the non-fertilized control soil and in MIN (Fig. 2C). The initial DOC concentrations in the control soil and in MIN were higher than the apparent lower limit for denitrification determined by Groffman et al. (1996) and Weymann et al. (2010). The recovery of the applied C was close to 100% in SUC treatment as the application rate was 500 mg C kg−1 dry soil, and measured DOC concentration after 6 h was 544 (±5.2) mg kg−1 dry soil in SUC. DOC concentrations decreased to the background level 120 h after start of the treatments in SUC and MST. The decrease in DOC concentration was more rapid in SUC than in MST which is most likely due to differences in the degradability of soluble C compounds, as well as to a gradual supply of new DOC due to hydrolysis of polymers present in the maize straw. In BGR, the DOC concentration was 349 (±20) mg C kg−1 dry soil 6 h after the start of treatments, and interestingly concentrations remained almost constant during the experiment. This confirms that BGR contains a large fraction of soluble but relatively recalcitrant organic compounds (Cayuela et al., 2010).

icant microbial NO3 − assimilation, estimated denitrification rates (nitrate reduction rates calculated from linear regression) were 13.5, 10.9, 5.3, 4.6, 2.8, 2.8 mg NO3 − -N kg−1 dry soil day−1 for MST + SUC, MST, SUC, BGR, MIN and control soil respectively. In addition, overall cumulative CO2 fluxes (Fig. 3) showed very high correlation (R2 = 0.9) with the net NO3 − loss (initial NO3 − concentration minus final). Pal et al. (2010) reported with their aerobic and anoxic incubation experiment that nitrate reduction rates due to denitrification were 3–56 mg NO3 − -N kg−1 dry soil day−1 depending on the quality of the available carbon in soil amended with peat. This report is in line with our findings that not only the quantity but also the quality (degradability) of the applied carbon compounds in soil are key parameters governing the denitrification

3.1.3. N2 O emissions For the non-fertilized control and MIN, timing, rates and cumulative N2 O emissions were almost identical (Fig. 1B). This suggests that the addition of mineral-N fertilizer alone (in the form of ammonium) did not enhance the formation of N2 O from either nitrification or denitrification throughout the incubation period. This is probably because of low nitrification rates at this soil moisture level (67% WFPS) as confirmed by stable NH4 + concentrations throughout the incubation period in MIN (see Section 3.1.1). Dobbie and Smith (2001) reported an exponential increase in the fraction of soil volume which is anaerobic due to either increasing soil respiration or decreasing oxygen diffusion or both with WFPS shares of 50% or more. With the exception of BGR, cumulative N2 O fluxes were

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Fig. 4. Time course of N2 O fluxes from soil (mg N2 O-N kg−1 dry soil h−1 ) and soil NO3 concerntrations (mg NO3 −1 -N kg−1 dry soil) after application of bigas residue (BGR; at an N rate of 150 kg NH4 -H ha−1 ), maize straw + mineral-N (MST; at an N rate of 150 kg NH4 + -N ha−1 ), sucrose + mineral N (SUC; at an N rate of 150 kg NH4 + -N ha−1 ), and maize straw + sucrose + mineral-N (MST + SUC; at an N rate of 150 kg NH4 + -N ha−1 ) during 124 h of the pot experiment (Exp. 1). Error bars show the standard error of the mean of each treatment (n = 3). In some cases error bars are smaller than the symbols.

significantly higher in soils with organic substrate treatments than in the control soil (Table 3). Greatest N2 O fluxes were measured in SUC. Maximum N2 O emission rates were found 24 h after start of the treatments, and fluxes sharply decreased to background levels 30 h after start of the experiment in MST and MST + SUC. In SUC however, N2 O fluxes decreased much later to background levels, i.e. only 65 h after start of treatments. Interestingly, the addition of biogas residue to the soil did not induce a very sharp N2 O peak as observed with the other organic substrates. The emission rates clearly went down after 65 h. Overall in the treatments with organic substrate addition, sharp reductions in N2 O emission rates occurred just when soil NO3 − concentrations fell below 20 mg NO3 − -N kg−1 dry soil (Fig. 4), while CO2 emission rates in the respective treatments were still very high at these moments. In addition, DOC concentrations were still sufficient to promote high denitrification rates (Fig. 2C). Therefore, we hypothesize that while denitrification was still going on, N2 O reduction became greater than the N2 O production when NO3 − concentrations fell below 20 mg NO3 − -N kg−1 dry soil in all soils amended with organic substrates i.e. the share of N2 O among denitrification products (N2 O/(N2 O + N2 )) declined. It has been observed in previous studies that the N2 O/(N2 O + N2 ) product ratio of denitrification was correlated with the ratio of NO3 − to available C in arable soils (Weier et al., 1993; Van Cleemput, 1998), in tropical forests (Kiese and Butterbach-Ball, 2002) and in laboratory incubation experiments (Alinsafi et al., 2008). If we attribute the decline in soil NO3 − concentrations to denitrification only (i.e. denitrification being the only NO3 − -N consuming process), the estimated share of N2 O/(N2 O + N2 ) would be close to 1 for all treatments in the initial period of the experiment. However, when NO3 − concentrations in soil fell below 20 mg NO3 − -N kg−1 dry soil, the ratio decreased sharply, ending up close to zero in organic substrate amended soil (Fig. 4). Therefore, we hypothesize that the effects of the ratio between NO3 − concentration and

available C on the share of N2 O on denitrification products can only be seen when the soil NO3 − concentration decreases under a threshold value (around 20 mg NO3 − -N kg−1 dry soil in this experiment). This data may suggest that a kind of site-specific threshold value may apply in N-fertilized soils that leads to high N2 O reduction thus lower N2 O emission. However, neither in this incubation experiment nor in the above mentioned reports, molecular N2 gas emission has been measured, therefore, the final proof for this conclusion is still missing. 3.2. Experiment 2 (robot experiment) To present the kinetics of NO, N2 O and N2 accumulation during the incubation and to illustrate the calculation for the N2 O product ratio of denitrification, data from incubation study are shown in Fig. 5. Kinetics of NO, N2 O and N2 accumulation during the incubation period were used for the calculation of the N2 O/(N2 O + N2 ) product ratio of denitrification and the total denitrification rate. Mean denitrification rates as well as N2 O/(N2 O + N2 ) ratios of denitrification were calculated for periods when concentrations of N2 and N2 O increased linearly. 3.2.1. Denitrification potential of soils Maximum soil denitrification rates were greater in sandy-loam soil (SL) than in the long-term organic matter-amended sandy soil (LOM), and lowest rates were observed in the long-term mineral N treated sandy soil (LMN) that had not received any organic fertilizers (Fig. 6A). As soil types were similar in LOM and LMN, the higher denitrification rate in LOM than LMN can be attributed to the difference in long-term organic carbon input. Denitrification rates in soils treated with 0.2 mM KNO3 were 0.96 mg N kg−1 dry soil day−1 in LMN and LOM and 0.72 mg N kg−1 dry soil day−1 in SL. With 2 mM KNO3 , denitrification rates in LOM and SL were higher than with 0.2 mM KNO3 (Fig. 6A). Here, denitrifi-

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Fig. 5. Time course of NO, N2 O and N2 concerntrations in closed incubations flasks as affected by N rates (0.2, 2, 10, 20 mmol KNO3 ) in long-term organic matter treated soil (LOM).

cation rates were 3.12, 2.16, and 0.96 mg N kg−1 dry soil day−1 in SL, LOM and LMN respectively. Denitrification was most likely limited by NO3 − in LOM and SL at N rates below 2 mM KNO3 (Fig. 5). A positive relation between denitrification rates and NO3 − concentration has been reported when the concentration of NO3 − was below 1 mM N in sediment slurry experiments (Zhong et al., 2010; Ogilvie et al., 1997). However, the lack of effects of the initial KNO3 concentration on denitrification potential of LMN suggests that the latter relation highly depends on the availability of carbon compounds in the soil. Maximum denitrification rates were measured at an N level of 2 mM KNO3 in LOM and SL. Surprisingly, denitrification rates showed a decreasing trend with increasing KNO3 concentrations from 2 mM KNO3 to 20 mM KNO3 . As far as we are aware, there are no previous reports showing the negative effect of higher NO3 − concentrations on the denitrification rate at such a moderate N level. A potential inhibitory effect of high NO3 − concentrations on the N2 O-reduction process will be discussed in Section 3.2.2. 3.2.2. N2 O product ratio of denitrification The N2 O/(N2 O + N2 ) product ratio of denitrification showed very high variation, it ranged between 0.02 and 0.94 when comparing all treatments (Fig. 6B). The lowest N2 O/(N2 O + N2 ) product ratios were observed at N concentrations of 0.2 mM NO3 − in all soils. Denitrification rates were limited by NO3 − availability in soils treated with 0.2 mM KNO3 especially in SL and LOM. In situations where the organisms experience a shortage of NOx , the relative rate of N2 O reduction (in relation to N2 O production) increases, thus very low N2 O/(N2 O + N2 ) product ratios of denitrification dominate. Based on a similar incubation experiment, Liu et al. (2010) reported that denitrifiers used an existing denitrification proteome from the beginning of anoxic treatments. In all soils, denitrification N2 O product ratios increased significantly with higher NO3 − concentrations. The N2 O product ratio of denitrification already reached values close to maximum at an N

concentration of 2 mM KNO3 in LMN. Interestingly, the N2 O product ratio of denitrification was clearly lower in LOM than in LMN at the two lower N levels which we attribute to the higher C availability and higher denitrification rate of LOM than LMN. Higher C availability in LOM promoted the flow of electrons driving denitrification which further favoured the last step of denitrification, e.g. the reduction of N2 O to N2 (Weymann et al., 2010). The maximum N2 O/(N2 + N2 O) product ratio of denitrification was found at 10 mM KNO3 in all soil types. Van Cleemput (1998) concluded that nitrate usually inhibits or retards N2 O reduction, resulting in higher N2 O release, but claimed that this type of inhibition would seldom occur at the typical nitrate concentrations found in soil. However, our results clearly indicate that 10 mM KNO3 is sufficient to minimise the last step of denitrification (N2 O to N2 ), and drastically increase N2 O formation in all soil types tested in this experiment (Fig. 6C). Importantly, such concentrations can easily be reached by normal mineral N fertilizer applications, and may remain at such levels for some time depending on the subsequent crop N uptake. Weier et al. (1993) concluded from their results that in moist or wet soils the addition or presence of an easily decomposable Csubstrate would increase the conversion of N2 O to N2 . The two incubation experiments presented here suggest that this is true for low nitrate concentrations, but at NO3 − concentrations ≥10 mM, this is no longer the case. This information may be used to set a kind of specific threshold value indicating the range of enhanced N2 O reduction and mitigation of N2 O emission by organic matter addition. This hypothesis however needs verification both in laboratory experiments and under field conditions. With soil SL and an N supply of 20 mM KNO3 , the N2 O product ratio of denitrification was 0.94 which was considerably higher than with the two other soils tested, LOM and LMN. So for soil SL the data indicate that, at these high NO3 − concentrations the denitrifier community almost ceased to reduce N2 O, thus they would almost exclusively produce N2 O when the NO3 − concentrations surpass a threshold. In the more sandy soils, LOM and LMN, the

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4. Conclusion In the present soil incubation experiments, it has been shown that application of organic fertilizers with large amounts of labile C induced drastic increases in the denitrification rate. NO3 − concentrations limited denitrification only in the range of very low concentrations (0.2 mM KNO3 treatment). However, in soils with a soil solution of more than 2 mM NO3 − , the denitrification rate decreased significantly. This negative relationship of NO3 − concentrations greater than 2 mM and denitrification rate was attributed to an inhibitory effect of NO3 − on N2 O reduction at such moderate N levels. Secondly, a clearly lower N2 O/(N2 O + N2 ) product ratio was found with long-term application of organic matter. By applying high concentrations of nitrate however, this effect disappeared completely because of the inhibitory effect of NO3 − on N2 O reduction. This implies that an increase in the soil organic matter content may lower the N2 O/(N2 O + N2 ) ratio, but only at low fertilizer levels. Thus in many soils, especially those with moderate or high N contents, the beneficial effect of organic matter amendment on greenhouse gas reduction by soil C sequestration may be compensated or even over-compensated by additional N2 O emissions. Our conclusions have been drawn from laboratory incubation experiments as methodology to study N2 O reduction in the field, e.g. studies of N2 O/(N2 O + N2 ) product ratios under field conditions are still missing. The study highlights the relevance of N2 O reduction in fertilized arable soils. Laboratory studies can contribute knowledge details on N2 O reduction in fertilized soils, however in future this must be complemented by field studies and appropriate analytical techniques.

Acknowledgements We are thankful to the Schleswig-Holstein Ministry of Economy and Science for funding part of this work in the scope of the BiogasExpert program. We would also like to acknowledge contributions by Dr. Joachim Lammel, Dr. Klaus Blankenau, and we thank two anonymous reviewers for their constructive comments.

References

Fig. 6. Effect of different KNO3 concentrations on soil N turnover and emission rates in long-term mineral N treated sandy loam soil (LMN), long-term organic manure treated sandy loam soil (LOM), and in Lavesum soil (SL): (A) denitrification rate, (B) share of N2 O in total denitrification (N2 O/(N2 O + N2 )) and (C) maximum N2 O emission rate in the robotized incubation some cases error bars are smaller than the symbols.

maximum N2 O production ratio of denitrification was only 0.68, indicating that N2 O reductase was still functioning, even at quite high NO3 − concentrations. Weier et al. (1993) reported that N2 O product ratio of denitrification was more than 0.90 in treatments with high NO3 − concentration although lower ratios were also reported. Therefore, data of our incubation experiment contribute to resolving this question as they clearly indicate that maximum N2 O product ratios of denitrification vary between soil types, and this relationship probably also relates to variations in the microbial communities in different soils. On the other hand, the lack of significant differences in N2 O production ratio between LOM and LMN at high NO3 − concentrations indicates that long-term application of organic fertilizer did not cause any significant change in the behaviour of denitrifiers in these sandy soils when soil NO3 − was above the threshold value.

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