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Journal of soil science and plant nutrition

On-line version ISSN 0718-9516

J. Soil Sci. Plant Nutr. vol.16 no.2 Temuco June 2016  Epub May 04, 2016

http://dx.doi.org/10.4067/S0718-95162016005000029 

Effect of Carbon source on dissimilatory nitrate reduction to ammonium in Costal Wetland sediments

X. Liu1,2, J.-G. Han3, Z.-W. Ma1,2, Q. Wang4, L.-H. Li1*

1State Key Laboratory of Desert and Oasis Ecology, Xinjiang Institute of Ecology and Geography, Chinese Academy of Sciences, Urumqi, 830011, Xinjiang, China.

*Corresponding author: lilh@ms.xjb.ac.cn

2University of Chinese Academy of Sciences, Beijing, 100049, China.

3College of Biology and the Environment, Nanjing Forestry University, Nanjing, 210037, China.

4Faculty of Agriculture, Shizuoka University, Shizuoka 422-8529, Japan.

 


Abstract

15N tracing technique was applied to investigate the effects of various organic carbon (OC) sources on dissimilatory nitrate (NO3-)reduction to ammonium (NH4+) (DNRA) rates in the coastal wetland sediments. Soils collected from the Chongming Dongtan wetland were incubated at 25oC in the dark for 24 h following the additions of OC sources (glucose, acetate, malate, citrate and oxalate (500 μg C g-1 dry soil)) and 15N-labeled NH415NO3 (initial 15N atom% of NO3--N is 20%). The results showed that soil DNRA rates varied from 0.018-0.497 mg N kg-1 dry soil d-1during the whole incubation, and the rates differed significantly among treatments following the order: oxalate> citrate> glucose> acetate> malate> no exogenous C addition over the first 12-h incubation. This was possibly caused by the different decomposition rates of various OC sources, which further influenced the available energy provided for DNRA microorganisms. Soils with no addition of exogenous C showed lower soil DNRA rates, presumably because of the low C/NO3- ratio as well as energy availability. The relative lower soil DNRA rates over the 24-h incubation indicated that DNRA is a fast process. Our results suggest that DNRA could be controlled by OC sources, especially organic acids, demonstrating that the widespread use of glucose in soil laboratory studies might lead to misleading knowledge on understanding the effects of OC on soil DNRA process.

Keywords: Dissimilatory nitrate reduction to ammonium, Carbon source, 15N tracing technique, organic acid, coastal wetland

 


1. Introduction

Nitrogen (N) is an essential nutrient affecting biological production in terrestrial and aquatic ecosystems, hence it is important to understand how this element is cycled in a specific region or ecosystem (Francis et al., 2007; Zhao et al., 2013). Coastal wetlands reside at a critical interface between terrestrial land and ocean, which plays a crucial role in balancing and sequestrating N as well as protecting biodiversity of coastal ecosystems (Bai et al., 2012). N has been believed to be a major limiting nutrient for plant growth in pristine coastal wetlands. However, an increase in human disturbance along the coast over recent decades has accelerated N loading into coastal areas via nutrient-rich rivers, groundwater as well as atmospheric N deposition (Koop-Jakobsen and Giblin, 2010). As a result, coastal ecosystems have always experienced eutrophication, which caused severe environmental problems, such as increased frequencies and magnitudes of phytoplankton blooms and loss of habitats for submerged vascular plants (Rabalais and Nixon, 2002). An improved understanding of N dynamics in coastal wetlands is needed to maintain regional and global ecological security (Bai et al., 2012). For a multitude of coastal wetlands, the dominant form of anthropogenic N inputs is nitrate (NO3-) (Burginand Hamilton, 2007). There are two important NO3- removal pathways in wetland sediments: one is denitrification (DNF), which sequentially reduces NO3- to dinitrogen (N2) via nitrite (NO2-), nitric oxide (NO) and nitrous oxide (N2O), and the other is dissimilatory NO3- reduction to ammonium (NH4+) (DNRA), by which NO3- is reduced to bioavailable NH4+ (Zumft, 1997; Rütting et al., 2011). Because of the different productions, DNF has been not only considered as a major pathway of N loss in soil-plant systems (Gong et al., 2013), but also the major biological process responsible for global increases in atmospheric N2O (Baggs, 2008; Morley et al., 2014). In contrast, DNRA conserves N in soils as NH4+, thus retaining N levels of soils (Yin et al., 2002; Koop-Jakobsen and Giblin, 2010). For a long time, DNF was considered as the predominant pathway to remove NO3- (Burginand Hamilton, 2007). However, over the last few decades, applications of 15N tracer technique and microbial molecular methods have generated growing evidence that DNRA is also an important N reduction process and cannot be ignored (Giblin et al., 2013). As an example, Giblin et al., (2013) compared the importance of DNF versus DNRA at 55 coastal sites and found that DNRA accounted for more than 30% of the NO3- reduction at 26 sites. Consequently, understanding what controls the DNRA process is vital to regulating N loadings in coastal wetland sediment.

Among various factors influencing DNRA carbon (C)/NO3- ratio has been confirmed to be one of the most important drivers regulating DNRA. For example, Tiedje (1988) argued that DNRA was favored when the ratio of labile C/NO3- was higher, and later studies further verified his hypothesis (Silver et al., 2001; Yin et al., 2001). In addition, as an essential electron donor of DNRA, organic C (OC) modifies the rate of DNRA by supporting respiration or fermentation that influences the population of DNRA bacteria. However, despite this requirement for OC, the regulation of DNRA by OC is poorly understood compared to other soil parameters such as pH, temperature, moisture and redox status (Rütting et al., 2011), and most existing publications only took glucose into consideration (Henderson et al., 2010; Lu et al., 2013). Nevertheless, plant-derived low molecular C compounds in soils, which are key factors affecting the rhizosphere microbial community, are always ignored (Morley and Baggs, 2010; Morley et al., 2014). Different OC substrates may also have different effects on the activity of enzyme (Morley and Baggs, 2010). Morley and Baggs (2010) proposed that OC may alter the fate of NO3- in soils on the basis of these hypotheses. In addition, >Bonin et al. (1999) pointed out that the theoretical energy of various C substrates was quite different, which probably also affect the DNRA process. Consequently, application of glucose alone would not provide enough oversight of the regulatory factors controlling DNRA. In order to advance understanding with a more realistic perspective, other C sources, especially the organic acids generated from root exudates and the decomposition of organic matter in plant residues, need to be considered to elucidate their unknown influences on DNRA. Under the background of rapid development of coastal economy, understanding the effects of OC on soil DNRA will help formulating appropriate strategies for the reduction of high N loadings in soils of coastal ecosystems.

In this study, 15N-labelled NO3--N tracer technique was applied to determine the effects of individual additions of different C sources (glucose, acetate, malate, citrate and oxalate) and no exogenous C addition (here termed as NC) on DNRA process in sediments of Chongming Dongtan wetland in Changjiang Estuary. The organic acid selected could be released in soils by rhizosphere fungi, root exudates and the decomposition of organic matter in plant residues.

2. Materials and Methods

2.1. Study area

Chongming Island (31o25′N- 31o38′N, 121o50′E-122o05′) is the third largest island in China and one of the largest alluvial islands in the world, which is formed by alluvial silt and soil from the upper reaches of Yangtze River (Li et al., 2010). On the eastern end of the island is the Chongming Dongtan wetland, which is the largest and youngest natural tidal flat in the Changjiang Estuary (Zhang et al., 2011). It has a northern subtropical ocean climate, with an average annual temperature of 15.3oC, while mean monthly temperature ranges from 2.7oC in January to 27.6oC in August. The mean annual precipitation is 1022 mm, with more than 70% of the annual rainfall occurring between April and September (Li et al.,2010). The vegetation area of the tidal flat covers 27.51 km2, and the vegetations primarily consists of Phragmites communis and Scirpus mariqueter. However, with its recent rapid spread and growth, the invasive Spartina alterniflora has occupied 33.1% of the vegetation area, and mainly distributed in the northeast of Chongming Dongtan wetland (Gan et al., 2009).

2.2. Soil sampling

Soil samples (0-15 cm depth) were collected from a site covered with Phragmites communis (31o37′N, 121o42′E) in Chongming Dongtan wetland in October 2013, and they were transported to the laboratory within 12 h. Once returned to the laboratory the soils were air-dried in the shade and sieved (≤ 2 mm). The visible plant materials in sieved soils were then removed. It was a silty clay loam (sand 44.18%, silt 55.56%, clay 0.26%) with a pH (H2O) of 8.65, soil organic carbon (SOC) of 8.39g C kg-1 and total nitrogen (TN), NH4+-N and NO3-- N of 1000.00, 6.20 and 3.40 mg kg-1, respectively.

2.3. Experimental set-up

Soil samples (30 g at 5% gravimetric water content) were packed into 250 mL Erlenmeyer flasks, and then 60 mL of deionized water were added into the flasks. The flasks were then capped with breathable films. Afterwards, the flasks were pre-incubated at 25oC in the dark for 7 days. As soon as the pre-incubation finished, different OC substrates, including glucose (GC), acetate (AT), malate (ML), citrate (CT) and oxalate (OL), were applied into the flasks and each treatment had 6 flasks. In addition, 6 flasks with >no exogenousC addition (NC) were also included as a control treatment. In order to make all samples the same liquid volume, 10 mL of OC solutions (19.79 mM GL, 59.38 mM sodium acetate, 29.69 mM sodium malate, 19.79 mM sodium citrate or 59.38 mM sodium oxalate) or 10 mL deionized water (NC) was added to each treatment. The final C added in soils was equivalent to 500 μg C g-1. After additions of C, 1 mL 15N-labeled ammonium nitrate (NH415NO3) (15.2 mM at 30.15 atom % excess 15N) was applied into each flask which resulted in a target 15N enrichment of 20% of soil NO3--N pool. All amendments were added to flasks by syringes through the breathable films. Moreover, 3 additional flasks without addition of C and 15N-labeled were also prepared to measure the contents of organic N (ON), NH4+-N and NO3--N after pre-incubation, which represented the initial N contents before the 15N tracer experiment. Therefore, there were total 39 flasks used in this study.

After all amendments were added, the flasks were immediately incubated in the dark at 25 oC. It is generally believed that DNRA is a fast process (Yin et al., 2002; Lu et al., 2013). In order to make our investigation comparable to previous ones, 12 h and 24 h were chosen as the incubation time, thus 3 flasks of each treatment were destructively sampled after 12 and 24 h incubations, respectively. To extract soil inorganic N, 80 mL of 1.88 M KCl was added to each flask. Because 60 mL deionized water and 10 mL C solutions had been added into each flask at the start of the pre-incubation, the final KCl concentration was about 1 M. The soil extracts were analyzed for the concentrations as well as 15N atom% of inorganic N (NH4+-N and NO3--N). The extracted soils were then washed with deionized water and oven-dried at 60oC for the determination of concentrations and 15N atom% of TN in the residuals, which was considered to be ON.

2.4. Soil measurements

Soil NH4+-N and NO3--N concentrations in the extracts were determined colorimetrically by ultraviolet spectrophotometer (UV-2550, Shimadzu Co., Ltd., Kyoto, Japan). Soil TN and ON were measured by semimicro Kjeldahl digestion selenium (Se), copper sulfate (CuSO4) and potassium sulfate (K2SO4) as catalysts. SOC was determined by concentrated sulfuric acid (H2SO4) potassium dichromate (K2Cr2O7) oxidation and subsequent titration. Soil pH was determined by a combination electrode in the 1:5 soils to water suspension (w:v). Soil water content was determined gravimetrically after subsamples were dried at 105oC to a constant weight. The diffusion method (Brooks et al., 1989) was used for extracting soil 15NH4+-N and 15NO3--N while oven-dried soil was directly used for ON-15N atom% determination. An automated C and N analyzer interfaced with an isotope ratio mass spectrometer (Flash EA-Delta V Advantage, Thermo-Fisher Scientific Co., Ltd., Massachusetts, USA) was used for determining N isotope ratios.

2.5. Calculations and statistical analyses

According to the procedures of Buresh et al.,(1978), Chen et al.,(1995) and Yin et al., (2002), DNRA rate (d, mg N kg-1 dry soil d-1) and relative potential of DNRA (d', %) were calculated using Equation 1 and Equation 2, respectively:

where [NH4]t1 and [NH4]t2 are the soil NH4+-N concentrations at time t1 and t2 (mg N kg-1 dry soil), respectively; ft1 and ft2 are the soil NH4+-15N atom% at time t1 and t2 (%), respectively; [ON]t1 and [ON]t2 are soil ON concentration at time t1 and t2 (mg N kg-1 dry soil), respectively; ft1' and ft2' are the soil ON-15N atom% at time t1 and t2 (%), respectively; t1 is the beginning time of incubation and t2 is the end time of incubation.

where [NH4]t1, [NH4]t2, [ON]t1, [ON]t2, ft1, ft2, ft1', ft2', t1 and t2 are the same as the variables in Equation 1; [NO3]t1 and [NO3]t2 are the soil NO3--N concentrations at time t1 and t2 (mg N kg-1 dry soil), respectively; ft1'' and ft2'' are the soil NO3--15N atom% at time t1 and t2 (%), respectively. It should be noted that at the beginning of the first incubation stage, natural abundance of 15N (0.366%) was applied to stand for the value of NH4+-15N atom% and ON-15N atom%, while the value of NO3--15N atom% was 20% which resulted from the additions of NH415NO3 and calculated using the Equation 3 as following:

where atom% is the value of 15N enrichment, S is the concentration of soil nitrate, 0.366% is the natural abundance of 15N, A is the concentration of nitrate added after OC additions, E is the abundance of 15N of added NO3--N (30.15%).

To test whether there were differences among treatments in concentrations and 15N atom% of various N fractions, soil DNRA rate as well as relative potential of soil DNRA, one-way analysis of variance (ANOVA) followed by LSD post hoc test was performed using SPSS 16.0 (SPSS Inc., Chicago, USA). The t test was used to examine the differences in soil indexes mentioned-above between 12-h and 24-h incubation. OriginPro 8.0 was used for performing figures (OriginLab, USA).

3. Results

3.1. Concentrations and 15N atom% of soil N pools

After pre-incubation, soil NH4+-N, NO3--N and ON were 10.50±0.42, 4.13±0.26 and 942.00±9.00 mg N kg-1 dry soil, respectively, and their kinetics over the 24-h tracer experiment were illustrated in Figure 1. Compared with NC, additions of OC significantly decreased soil NH4+-N during both stages (P<0.05) (Figure 1-a). Different soil NH4+-N decrements were also detected among five OC treatments and the order was OL>ML>GC>AT>CT. In addition, soil NH4+-N concentrations were significantly lower over the 24-h incubation than that over the 12-h incubation in all treatments (P<0.05). OC treatments also had negative effects on soil NO3--N concentrations, and the lowest value occurred in treatment OL (Figure 1-b). As the incubation went on, soil NO3--N concentrations decreased in spite of the different treatments. Soil ON also decreased under different OC treatments, and showed the same response with inorganic N (Figure 1-c). Over the first 12-h incubation, soil ON concentrations followed the order: NC>AT>OL>GC>ML>CT, while the order changed to OL>ML>CT>AT>NC>GC over the second 12-h incubation. For treatments GC, AT, CT and NC, soil ON concentrations over the 24-h incubation were significantly lower in comparison with that of over the 12-h incubation (P<0.05). Nevertheless, significant differences were not observed in treatments ML and OL (P>0.05).

Figure 1. The concentrations of soil NH4+-N (a), NO3--N (b) and ON (c) in sediments of Chongming Dongtan wetland supplied with different forms of OC substrates or none (NC) after 12 and 24-h of incubations. The same uppercase letter indicates values are not significantly different among treatments (P<0.05) within a same incubation period. The same lowercase letter indicates values are not significantly different between two periods (P<0.05) with the same treatment. The error bars stand for standard deviations of the means (n=3).

As shown in Table 1, soil NH4+-15N atom% increased to 0.470-0.654% after 12-h incubation, and the highest and lowest NH4+-15N atom% were found in treatment OL and NC, respectively. The second 12-h incubation further increased NH4+-15N atom% in both treatments, although the values did not differ significantly in OC treatments between the two incubation stages (P>0.05). Soil NH4+-15N atom% of OL were significantly higher than those in other treatments (P<0.05). Soil ON-15N atom% also showed an increasing trend, although the increments were lower compared with the natural abundance of ON-15N atom%. During the entire incubation, soil ON-15N atom% ranged between 0.373% and 0.407%. On the contrary, significant reductions of 15N atom% were detected in soil NO3--N pool. After 24-h incubation, soil NO3--15N atom% decreased to just 0.570-1.270% in treatments with OC additions, which were lower compared with the initial NO3--15N atom%. However, these huge reductions did not occur in treatment NC. Furthermore, the NO3--15N atom% also differed significantly between the two incubation stages (P<0.05), indicating soil NO3--15N pool was continuously diluted over time.

Table 1. 15N atom% of various N fractions in the sediments of Chongming Dongtan wetland supplied with different forms of OC substrates or none (NC) after 12 and 24 h of incubations.

a Values are means (SD) (n=3).

b The same uppercase letter indicates values are not significantly different among treatments (P<0.05) within a same incubation time. The same lowercase letter indicates values are not significantly different between two periods (P<0.05) with the same treatment

3.2. Soil DNRA rate

Different OC sources significantly affected soil DNRA rates in both incubation stages (Figure 2). Soil DNRA rates varied from 0.018-0.497 mg N kg-1 dry soil d-1, and the maximum and minimum were measured in OL over the 12-h incubation and in AT over the 24-h incubation, respectively. For the first 12-h incubation, the order of DNRA rates followed: OL>CT>GC>AT>ML>NC, while soil DNRA rates in GC and AT did not differ significantly (P>0.05). Notably, the mean DNRA rate in soils of NC was only 0.033 mg N kg-1 dry soil d-1, which was significantly lower than that in soils with OC additions (P<0.05). Incubation time also influenced soil DNRA rates: soil DNRA rates over the second 12-h incubation were significantly lower than that over the 12-h incubation in both treatments except NC (P<0.05). During the second incubation stage, soil DNRA rates in treatment OL were not the highest due to their rapid decreases. Instead, the maximum of soil DNRA rates was detected in treatment AT, although its mean value was only 0.074 mg N kg-1 dry soil d-1.

Figure 2. DNRA rates in sediments of Chongming Dongtan wetland supplied with different forms of OC sources or none (NC) over 12 and 24 h of incubations. The same uppercase letter indicates values are not significantly different among treatments (P<0.05) within a same incubation period. The same lowercase letter indicates values are not significantly different between two periods (P<0.05) with the same treatment. The error bars stand for standard deviations of the means (n=3)

3.3. Relative potential of soil DNRA

Relative potential of DNRA, which also refers to the contribution of DNRA to total dissimilatory NO3- reduction, was estimated by the recovery of 15N in NH4+-N and ON divided by the changes of 15N in NO3--N. As shown in Figure 3, different OC treatments influenced the relative potential of soil DNRA during both incubation stages. For the first 12-h incubation, the relative potential of soil DNRA varied from 4.50-27.57%, and values from soils amended with CT were significant higher than other treatments (P<0.05), although its soil DNRA rate was not the highest (Figures 2 and 3). Moreover, similar to the DNRA rate, relative potential of soil DNRA also showed a decreasing tendency in both treatments during the second 12-h incubation.

Figure 3. Relative potential of soil DNRA in sediments of Chongming Dongtan wetland supplied with different forms of OC substrates or none (NC) after 12 and 24 h of incubations. The same uppercase letter indicates values are not significantly different among treatments (P<0.05) within a same incubation period. The same lowercase letter indicates values are not significantly different between two periods (P<0.05) with the same treatment. The error bars stand for standard deviations of the means (n=3)

4. Discussion

In this study, inorganic N (NH4+-N and NO3--N) and ON concentrations as well as the 15N atom% of NO3--N decreased over the course of the incubation, while 15N atom% of NH4+-N and ON increased with time. The characteristics of soil N concentrations along time were in line with some previous investigations (Yin et al., 2002; Scott et al.,2008; Lu et al.,2015). It was widely believed that NH4+-N is an important product of DNRA (Rütting et al., 2011; Giblin et al., 2013), and the NH4+-N generated from DNRA could be converted into other N pools through other processes, one of which is NH4+-N immobilization (Tobias et al., 2001; Yin et al., 2002; Zhang et al., 2014). For instance, Tobias et al., (2001) and Yin et al., (2002) claimed that NH4+-N generated from DNRA was mainly immobilized in the soil organic nitrogen pool, while Scott et al.,(2008) pointed out that DNRA provided little NH4+-N to nitrification. As the microorganisms were stimulated by exogenous C sources, NH4+-N immobilization might be enhanced, which would accelerate the consumption of NH4+-N (Yin et al., 2002). Therefore, NH4+-N in soils amended with OC decreased rapidly in contrast with that in NC. In addition, as a major reactant of DNRA as well as other NO3- reduction processes, NO3--N was continuously consumed and transformed into other N pools such as NH4+-N, ON, N2 and N2O. Different OC sources also affected NO3--N consumption rate, possibly through influencing activities of enzyme and/or microorganisms (Bonin et al.,1999). However, the NO3--N concentrations did not decrease drastically while the atom% of NO3--15N did, indicating that nitrification was occurring during incubations, which generated NO3--14N and diluted the NO3--15N pool. Furthermore, Rice and Tiedje (1989) pointed out that NO3- assimilation could be inhibited by 60% and 80% under low (0.1 mg N kg-1 dry soil) and higher soil NH4+- N concentrations (10 mg N kg-1 dry soil), respectively. In this study, NH4+-N concentrations in most treatments were higher than 10 mg N kg-1 dry soil over the 24-h incubations, hence NO3- assimilation was probably inhibited and the formed ON-15N might come from NH4+-N immobilization. The low NH4+-15N atom% also indicated that NH4+-15N generated through DNRA was probably converted into the ON pool.

Different OC additions significantly increased soil DNRA rates during the first 12-h incubation in the present study. As a common exogenous OC source, glucose has been widely selected to estimate the effect of C on soil DNRA process, whereas the existing results of its effect on DNRA are contradictory. Some studies demonstrated that the addition of glucose could support respiration as well as fermentation, thus stimulating DNRA (Buresh and Patrick, 1978; Yin et al.,2002). However, DeCatanzaro et al. ,(1987) and Chen et al., (1995) found that the addition of glucose did not influence soil DNRA. Some other publications also indicated that other C sources such as straw, glycerol, methanol and succinate had little effects on soil DNRA process (Buresh and Patrick, 1978; Yin et al.,1998). Buresh and Patrick (1978) and Yin et al., (1998) attributed these findings to the fact that the mentioned C sources were poor substrates for fermentation. However, there are three distinct pathways of microbial DNRA when available reductants such as sulfide and reduced iron are present:, they are fermentative DNRA, respiratory DNRA and Chemoautotrophic DNRA (Roberts et al., 2014). Hence, it is possible that the above mentioned C sources did not favor the three DNRA pathways controlled by the species of soil bacteria. In the present study, additions of glucose significantly increased soil DNRA rates compared with the NC treatment, indicating that bacteria involved in the DNRA process was exist in the collected sediments of Chongming Dongtan wetland. Nevertheless, which DNRA bacteria dominated the process are still unclear and deserve further investigations.

Low molecular organic acids are commonly present and are an important factor in affecting transformation and transfer of heavy metals in terrestrial ecosystem s (Schwab et al., 2008), although their effects on soil DNRA are still unknown. Results of this study indicate that the four organic acid treatments have positive effects on soil DNRA, and the effects on enhancing soil DNRA rate follows the order: OL>CT>AT>ML. The differences are presumably caused by the different decomposition rates of organic acids. In general, organic acids are decomposed by microorganisms through two pathways, oxidation and decarboxylation (McCollo et al., 2003). Some studies indicated that the decomposition rates of citrate and oxalate were faster than acetate (Jones et al.,2001; Hees et al.,2002). The fast degradations could rapidly provide energy for soil DNRA (Morley and Baggs, 2010) When the organic acids were almost entirely decomposed by soil microorganisms, the decomposition rates would be stable (Jones, 1998). Consequently, the soil DNRA rates in OL and CT were significantly higher than that in AT over the 12-h incubation, while they were higher in AT rather than OL and CT over the second 12-h incubation. Nevertheless, whether soil DNRA rate is associated to the decomposition rate of OC is still unclear and need further study. In the Chongming Dongtan wetland, the proportion of vegetation coverage area to total area is high (Gan et al., 2009). Large root exudates and plant residues may lead to high concentrations of soil organic acids. Consequently, glucose may not be the most appropriate C source for understanding the regulation of soil DNRA capacity in situ.

In treatment NC, the relative potential of DNRA were about 5%,which was in agreement with some previous studies (Tobias et al., 2001; Yin et al., 2002), implying that the SOC was limiting for soil DNRA, possibly due to lack of energy. Moreover, additions of 15N increased concentrations of soil NO3--N as well, which created low C/NO3- ratios in NC soils. It is widely accepted that the low C/NO3- ratio is unfavorable for DNRA (Rütting et al., 2011), thus the DNRA process must be weak in NC treatment. However, the effects of C/NO3- with different OC sources on soil DNRA are still unclear and need to be investigated in the future. In addition, soil NH4+-15N and ON-15N atom% showed little increase over the second 12-h incubation, and thus induced low soil DNRA rates in both treatments. This result demonstrated that DNRA was a fast process in soils, which was in consistent with some previous studies (Yin et al.,2002; Lu et al., 2013).

5. Conclusion

In this study, the effects of various C sources on soil DNRA rate were estimated using the 15N tracer technique. After the 24 h incubation, the results show that concentrations of inorganic N, ON as well as the 15N atom% of NO3--N have all decreased compared with the initial values,while 15N atom% of NH4+-N and ON showed increasing tendencies, indicating that bothDNRA, DNF, inorganic N immobilization and ON mineralization occurred during the incubation periods. Over the first 12-h incubation, soils with OL and CT exhibited greater soil DNRA rates as well as the relative potential of soil DNRA than that with GC, demonstrating that organic acids have different impacts on soil DNRA from glucose. The reason may be the different decomposition rates of various OC compounds influencing the available energy provided for DNRA microorganisms. In contrast, soils with no addition of exogenous C showed lower soil DNRA rates, presumably due to the lack of available energy and the low C/NO3- ratio. In addition, soil DNRA rates over the 12-h incubation were significantly higher than those over the 24-h incubation, demonstrating again that DNRA is a fast process. These findings indicate that soil DNRA can be regulated by different OC sources, suggesting that only taking glucose as C source may not reflect the soil DNRA capacity well in the field. In the future, more attention should be paid to investigate how soil pH, different quantities of various OC types as well as C/NO3- ratio with different OC types affect soil DNRA process.

Acknowledgements

This study was financially supported by the National Natural Science Foundation of China (41375149).

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