Urbanisation is increasing globally and China's land surface in particular is undergoing a transition from rural to urban landscapes (Zhu et al. Reference Zhu, Ioannidis, Li, Jones and Martin2011). Under this rapid and dynamic urbanisation, intensively anthropogenic activities can lead to elevated contents of common heavy metals (lead, copper (Cu), zinc and cadmium) in urban soils compared to rural soils. In addition to natural geogenic sources, the increasing heavy metal concentrations in urban soils are mainly derived from the atmospheric deposition of coal combustion (Luo et al. Reference Luo, Xue, Wang, Cang, Xu and Ding2015), followed by vehicle emissions. Among these so-called ‘urban-metals', Cu could be the most critical one with respect to the microbial denitrification process through which nitrate (NO3–) can be sequentially reduced to gaseous nitrous oxide (N2O) and dinitrogen (N2) via nitrite (NO2–) and nitric oxide (NO). The intermediate N2O is a powerful greenhouse gas with approximately 300-fold greater warming potential than CO2 and also contributes to stratospheric ozone destruction (Ravishankara et al. Reference Ravishankara, Daniel and Portmann2009). In particular, denitrification is the dominant nitrogen (N)-loss pathway in the terrestrial biosphere, especially in urban soils where the primary form of N-input is NO3– (Zhu & Carreiro Reference Zhu and Carreiro2004; Pickett et al. Reference Pickett, Cadenasso, Grove, Boone, Groffman, Irwin, Kaushal, Marshall, McGrath, Nilon, Pouyat, Szlavecz, Troy and Warren2011). Nitrate represents the most important substrate and electron acceptor for microbial denitrification. Moreover, Cu is known to function as an essential micronutrient, and, notably, the Cu-dependent nitrite reductase that converts non-gaseous to gaseous N as well as the N2O reductase that destroys N2O require Cu as a redox active metal cofactor (Felgate et al. Reference Felgate, Giannopoulos, Sullivan, Gates, Clarke, Baggs, Rowley and Richardson2012). Based on this, we expect a potential link between the soil Cu content and urban soil microbial denitrification.
For over three decades, Cu has been shown to be able to strikingly affect the activities of denitrifying reductases and to regulate the N2O emission from denitrification (Matsubara et al. Reference Matsubara, Frunzke and Zumft1982; Bardgett et al. Reference Bardgett, Speir, Ross, Yeates and Kettles1994; Sakadevan et al. Reference Sakadevan, Zheng and Bavor1999; Holtan-Hartwig et al. Reference Holtan-Hartwig, Bechmann, Høyås, Linjordet and Bakken2002; Granger & Ward Reference Granger and Ward2003; Magalhães et al. Reference Magalhães, Machado, Matos and Bordalo2011; Felgate et al. Reference Felgate, Giannopoulos, Sullivan, Gates, Clarke, Baggs, Rowley and Richardson2012; Moffett et al. Reference Moffett, Tuit and Ward2012; Sullivan et al. Reference Sullivan, Gates, Appia-Ayme, Rowley and Richardson2013; Black et al. Reference Black, Hsu, Hamonts, Clough and Condron2016). So far, quantitative DNA-based studies documenting how Cu impacts denitrification have been mainly conducted in microbial cultures (Felgate et al. Reference Felgate, Giannopoulos, Sullivan, Gates, Clarke, Baggs, Rowley and Richardson2012; Sullivan et al. Reference Sullivan, Gates, Appia-Ayme, Rowley and Richardson2013; Black et al. Reference Black, Hsu, Hamonts, Clough and Condron2016) and sediments (Magalhães et al. Reference Magalhães, Machado, Matos and Bordalo2011). Yet, very little is known about how Cu influences denitrification in soil environments at the gene level. Evidently, the potential effects have rarely been documented in urban soils, which are created by urbanisation and differ markedly from natural soils (Brown et al. Reference Brown, Johnson, Loveland and Theobald2005; Lehmann & Stahr Reference Lehmann and Stahr2007). Given that urban soils could receive nitrate and accumulate Cu due to urbanisation, Cu may profoundly affect urban soil denitrification. Genes encoding structurally different, yet functionally equivalent, nitrite reductases (nirK and nirS) and nitrous oxide reductase (nosZ) are commonly used to quantify different groups of denitrifying communities (Zumft Reference Zumft1997; Bru et al. Reference Bru, Ramette, Saby, Dequiedt, Ranjard, Jolivet, Arrouays and Philippot2011). Hence, this study evaluates the effects of Cu, a typical marker of urbanisation, on the abundance of nirK, nirS and nosZ genes in urban soils, which is a largely overlooked site but important hotspot for denitrification. Indeed, previous investigations exploring linkages between Cu and soil denitrification activity and N2O emission mainly used considerably high Cu amendments that generally exceeded 80μg Cu g–1 dry soil (Bardgett et al. Reference Bardgett, Speir, Ross, Yeates and Kettles1994; Holtan-Hartwig et al. Reference Holtan-Hartwig, Bechmann, Høyås, Linjordet and Bakken2002; Wang et al. Reference Wang, Burger, Doane, Horwath, Castillo and Mitloehner2013). Here we focus on lower Cu concentrations at levels typical for urbanisation to obtain conclusive data at realistic environmental conditions.
Thus, the current study aims to: (1) assess the denitrification function indicated by N2O emission and potential denitrification activity in a long-term undisturbed urban soil with distinct Cu amendments; and (2) investigate how Cu concentration affects denitrifier community abundance by detecting denitrification-related functional genes.
1. Materials and methods
1.1. Sampling
The topsoil (0–15cm in depth) was collected from Xiamen Botanical Garden [24°27′N, 118°05′E], China. The local climate is subtropical oceanic, with an average annual temperature of 21°C and a mean annual precipitation of 1200mm. The construction of the botanical garden began in 1960. Vegetation in the sampling site consists of trees with little grass. The selected soil has not been disturbed since 1989. After sampling, the urban garden soil was sieved (<2mm), then stored at 4°C prior to incubation and partially air-dried for chemical analysis. The soil had a texture of 73.54% sand, 25.22% silt and 1.24% clay, a pH of 5.8 and contained 28.3gkg–1 total carbon, 2.1gkg–1 total N and 0.4gkg–1 total sulphur.
1.2. Microcosm incubation
Soil samples were uniformly sprayed with a gradient of CuCl2 levels in a plastic valve bag and homogenised by hand-shaking for 30min. The concentrations of added CuCl2 were 0, 10, 20, 40, 80 and 160μg Cu g–1 soil (denoted as Cu0, Cu10, Cu20, Cu40, Cu80 and Cu160, respectively). After Cu exposure at 25°C in the dark for three days, the available total Cu concentrations were quantified, with values of 0.8, 7.4, 15.5, 29.5, 65.5 and 134.4μg Cu g–1 soil, respectively.
Potential denitrification activity was measured using the acetylene inhibition method (Myrold & Tiedje Reference Myrold and Tiedje1985). Soils equal to 10g of dry weight were transferred to 120ml serum bottles, and 5ml KNO3 solution (50mg N kg–1) was added as denitrification substrate. A small volume of deionised water was supplemented to attain a uniform moisture of 60% (w/w, absolute water content). For each treatment, three replicates were sampled to determine the initial nitrate concentrations, while the other nine replicates were sealed with butyl rubber stoppers and aluminium caps. All nine flasks were then evacuated and flushed with 99.999% helium for four cycles and finally vented to atmospheric pressure. Three of the anoxic flasks were injected with C2H2 (10% v/v of headspace) to inhibit nitrification and the reduction of N2O to N2 in denitrification. Immediately afterwards, all anoxic flasks were instantly incubated at 25°C in the dark for 21h using a robotised incubation system coupled to a gas chromatograph (see Section 1.3), as previously described (Molstad et al. Reference Molstad, Dörsch and Bakken2007). The moment of acetylene injection was regarded as the start of incubation (denoted as 0h). The flasks with C2H2 and three of the replicates without C2H2 were used to determine N2O and CO2 concentrations in the headspace at 0h and subsequently at 5, 9, 13, 17 and 21h of incubation; the remaining three replicates without C2H2 were only used for DNA extraction after 21h incubation. Soils in flasks without C2H2, which were used for gas monitoring, were collected to determine NO3– concentrations at the end of incubation.
1.3. Soil and gas measurements
The total Cu concentration in soil was analysed using a strong acid (HNO3–HClO4) digestion method (Lee et al. Reference Lee, Li, Shi, Cheung and Thornton2006), and determined by inductively coupled plasma-optic emission spectroscopy (ICP-OES) (Optima 7000 DV, PerkinElmer, USA). The available Cu in soil was extracted in 0.1M HCl (Aoyama & Nagumo Reference Aoyama and Nagumo1996; Xu et al. Reference Xu, Zhou, Zhan, Gao, Dou and Sun2013) and assessed by atomic absorption spectroscopy (M6, Thermo Fisher, USA). Soil pH was determined with a dual-channel pH-ion-conductivity-dissolved oxygen meter (X60, Fisher Scientific) using a soil–water suspension (1:2.5 w/v). Particle size distribution was analysed using a laser particle size analyser (Malvern Mastersizer 2000, UK). Total carbon, N and sulphur were determined using an element analyser (Vario MAX CNS, Germany). Soil NO3– was extracted in deionised water (Fang et al. Reference Fang, Koba, Makabe, Zhu, Fan, Liu and Yoh2012) and quantified by ion chromatography (Dionex ICS-3000, USA). Concentrations of headspace N2O and CO2 were analysed with an Agilent 7890 gas chromatograph (Santa Clara, CA, US). The robotised incubation system (Molstad et al. Reference Molstad, Dörsch and Bakken2007) was coupled to this gas chromatograph through an autosampler, allowing the direct and sequential measurement of headspace gases.
1.4. DNA extraction and quantitative polymerase chain reaction (qPCR)
Soil DNA was extracted from approximately 0.5g (wet weight) microcosm samples using the FastDNA SPIN Kit for soil (MP Biomedicals, Santa Ana, CA, USA) following the manufacturer's instructions. The purity and concentration of DNA were checked using an ultraviolet–visible spectroscopy spectrophotometer ND-1000 (NanoDrop, USA). The DNA was then stored at −20°C until real-time qPCR analysis.
Quantitative PCR amplifications were performed on a Light-Cycler Roche 480 instrument (Roche Molecular Systems). The bacterial 16S rRNA gene and key denitrification functional genes nirK, nirS and nosZ were amplified and quantified using the primer pairs 341F/517R (Muyzer et al. Reference Muyzer, De Waal and Uitterlinden1993), FlaCu/R3Cu (Hallin & Lindgren Reference Hallin and Lindgren1999), cd3aF/R3cd (Throbäck et al. Reference Throbäck, Enwall, Jarvis and Hallin2004) and nosZ-F/nosZ-1622R (Throbäck et al. Reference Throbäck, Enwall, Jarvis and Hallin2004), respectively. Twenty microlitres of PCR reaction volume contained 2μL of DNA (1–10ng), 0.25μM of each primer, 10μL of SYBR Premix Ex TaqII (TaKaRa) and 0.2μL of 0.1% bovine serum albumin (BSA) (TaKaRa). For the bacterial 16S rRNA gene, the thermal cycling condition was 95°C for 3min, followed by 40 cycles of 95°C for 15s and 56°C for 34s. For the nirK and nirS genes, the thermal profile was 95°C for 2min, followed by 40 cycles of 95°C for 30s, 57°C for 1min and 72°C for 1min. For the nosZ gene, the condition was 95°C for 1min, followed by 40 cycles of 95°C for 15s, 60°C for 15s and 72°C for 34s. The PCR reaction mixtures without template DNA were used as negative controls. All real-time PCR standard curves were generated from gradient dilutions of plasmids with known copy numbers. The amplification efficiencies for all assays were above 90%, with R 2 values of more than 0.99.
1.5. Statistical analysis
All experimental results are given on an oven-dry basis unless otherwise noted. Statistical analysis was conducted with SPSS software 18.0 for Windows. One-way ANOVA with Duncan's test was performed to determine the statistical significance of all data between treatments at P<0.05. Pearson correlation analysis was used to evaluate the relationship among independent variables at P<0.05 or P<0.01.
2. Results
2.1. Potential denitrification activity and N2O emission
The nitrate consumption rate decreased with increasing Cu amendments, and dropped to the lowest level (0.041μmol NO3– g–1 soil h–1) in the Cu160 treatment (Fig. 1a). The nitrate consumption rate in Cu-treated soils was found to be significantly lower (P<0.05) than that in the control until the Cu addition reached the value of 40μg Cu g–1 soil.
Figure 1 Mean (± standard error; n=3) nitrate consumption rate (a), potential denitrification activity (b), N2O emission from denitrification (c) and anoxic respiration rate (d) in soil microcosms with different Cu additions through the entire incubation. Matching letters above the bars indicate treatments without significant differences (P<0.05).
The N2O emission in treatments with C2H2 represented the potential denitrification activity, due to the inhibitory properties of C2H2 on nitrification and N2O reduction. In contrast, the N2O release in C2H2-free treatments represented the N2O emission via denitrification. We observed that the potential denitrification activity and denitrification-derived N2O emission decreased progressively with evaluated Cu concentrations and differed significantly among all treatments. As shown in Figure 1b, in the highest Cu-amended soil (Cu160 treatment), the potential denitrification activity (2.74nmol N2O g–1 soil h–1) was significantly (P<0.05) reduced by 40% compared with the non-amended control, and the corresponding N2O emission of denitrification (2.49nmol N2O g–1 soil h–1) decreased by 44% (Fig. 1c). In general, the N2O/(N2O+N2) ratios of denitrification products ranged from 91 to 96%, indicating that the dominant product of denitrification in the investigated soil was N2O rather than N2.
Anoxic respiration (CO2 emission rate) was determined from soils incubated without C2H2, since C2H2 might serve as a carbon source for soil microbes. As shown in Figure 1d, the Cu additions significantly (P<0.05) inhibited the CO2 emission rate with increasing concentrations. The respiration rate was 27.55nmol CO2 g–1 soil h–1 in the Cu160 treatment – 22% lower than that in the non-amended microcosm.
2.2. Abundance of denitrifying functional genes
The Cu additions exerted no significant effect on general bacterial abundance (quantified by 16S rRNA gene) with the exception of the Cu160 treatment, in which the abundance (1.03×1010 copies g–1 soil) was significantly (P<0.05) reduced by 55% compared to the control (Fig. 2a).
Figure 2 Mean (± standard error; n=3) bacterial 16S rRNA gene (a), nirK gene (b), nirS gene (c) and nosZ gene (d) abundance in soil microcosms with different Cu additions through the entire incubation. Matching letters above the bars indicate treatments without significant differences (P<0.05).
Further on, our results showed that Cu decreased the abundance of nirK and nirS genes with increasing concentrations. The significant difference on nirK gene abundance was only observed in the Cu160 treatment (8.83×107 copies g–1 soil) – 52% lower than that in the control (Fig. 2b). The abundance of the nirS gene in amended microcosms was not significantly affected compared to the control until the Cu concentration was as high as 40μg Cu g–1 soil (Fig. 2c). In the Cu40 treatment, copy numbers of the nirS gene (2.56×107 copies g–1 soil) were reduced by 26% compared to the control. In the Cu160 treatment, the nirS gene abundance decreased to 1.59×107 copies g–1 soil – 54% lower than that in the non-amended control. No significant difference was found for the nosZ gene copies across all treatments (Fig. 2d).
Of the denitrification functional genes analysed in this study, the abundance of the nirK gene in the control and amended microcosms were one order of magnitude larger than that of nirS and nosZ genes in respective microcosms.
2.3. Correlations between denitrification function and denitrifier abundance
The relationship between Cu additions, geochemical parameters indicating denitrification function, as well as denitrifier abundance, were examined using Pearson correlation analysis. As shown in Table 1, the Cu additions exhibited a strong and negative correlation (P<0.01) with respectively potential denitrification activity, N2O emission of denitrification, anoxic respiration and the abundance of nirK and nirS genes. On the other hand, there was no significant correlation with nosZ gene abundance. Nitrate consumption rate and bacterial 16S rRNA gene copies were negatively correlated (P<0.05) to Cu additions. Of the correlation between denitrification function (including both potential denitrification activity and denitrification-derived N2O emission) and denitrifying gene abundance examined, the potential denitrification activity showed a clear and positive correlation with nirK (P<0.01), nirS (P<0.01) and bacterial 16S rRNA gene copies (P<0.05), yet was not found to be correlated to nosZ gene abundance.
Table 1 Pearson correlations between denitrification function and denitrifier abundance. Abbreviations: Cu=copper additions; NO3– = nitrate consumption rate; PDA=potential denitrification activity; D-N2O=N2O emission from denitrification; D-N2=N2 emission from denitrification; CO2=anoxic respiration rate; NS=not significant at the 0.05 level (two-tailed).
1 Correlation is significant at the 0.05 level (two-tailed).
2 Correlation is significant at the 0.01 level (two-tailed).
3. Discussion
3.1. Effects of Cu on soil denitrification
Soil denitrification functions can be reflected by the potential denitrification activity and the N2O emission from denitrification. Our study showed that Cu additions significantly inhibited potential denitrification activity and denitrification-derived N2O emission, even at the lowest amendment (10μg Cu g–1 soil). The observed inhibitory effects of Cu on denitrification function have also been described in soils and sediments in previous studies (Bardgett et al. Reference Bardgett, Speir, Ross, Yeates and Kettles1994; Sakadevan et al. Reference Sakadevan, Zheng and Bavor1999; Holtan-Hartwig et al. Reference Holtan-Hartwig, Bechmann, Høyås, Linjordet and Bakken2002; Wang et al. Reference Wang, Burger, Doane, Horwath, Castillo and Mitloehner2013), where higher concentrations generally exceeding 80μg Cu g–1 dry soil or 500μg Cu g–1 dry sediment were required. In this investigation, Cu amendments exhibited suppression on denitrification functions at a level as low as 10μg Cu g–1 soil. Similarly, Magalhães et al. (Reference Magalhães, Machado, Matos and Bordalo2011) reported that the potential denitrification rate in an estuarine sediment were inhibited when amended with 4μg Cu per gram of wet sediment. However, we observed that the Cu additions had no significant effects on the N2 emission rate, as calculated by subtracting denitrification-derived N2O emission from corresponding potential denitrification activity in respective treatments (data not shown). Taking into account that N2O was the prominent denitrification product in this urban park soil across all treatments, especially in non-amended microcosms, the dominant denitrification pathways in our study were the precedent reactions generating N2O rather than the final step that reduces N2O to N2, and Cu decreased denitrification mainly by inhibiting the sequential reduction of NO3– to N2O. Increased Cu levels were also observed to inhibit overall microbial activity, as indicated by the decreased soil respiration. Additionally, our study revealed that different reduction steps involved in denitrification differed in sensitivity to Cu amendments. By comparing responses of nitrate consumption and denitrification-derived N2O emission to Cu additions, we at least can conclude that the nitrate reduction step was more tolerant than the entire two subsequent reduction steps by which NO2– was reduced to N2O via NO. Of particular interest was the observation that Cu additions did not impact N2 emission rate or stimulate N2O accumulation, indicating that the N2O reduction step was quite tolerant to Cu additions. This is not in line with the results of other studies (Holtan-Hartwig et al. Reference Holtan-Hartwig, Bechmann, Høyås, Linjordet and Bakken2002; Magalhães et al. Reference Magalhães, Costa, Teixeira and Bordalo2007, Reference Magalhães, Machado, Matos and Bordalo2011), suggesting that the N2O reduction step was highly inhibited by Cu amendments, and thereby enhanced the N2O emission from denitrification.
Our results showed that the copy numbers of the nirK gene detected were as high as 108 copies g–1 soil, while the abundance of nirS and nosZ genes were around 107 copies g–1 soil, respectively. Furthermore, across all treatments, the nirS gene was more abundant than the nosZ gene. The predominance of nirK-bearing denitrifiers and the greater abundance of nirS-bearing denitrifiers than nosZ-type denitrifiers could support our observation that the precedent reactions generating N2O were the dominant denitrification pathways in this soil rather than the N2O reduction. This was further supported by the significant and positive correlations between the abundance of nirK and nirS genes and the N2O emission from denitrification, indicating that, also, the denitrifying-associated genes abundance could effectively predict the corresponding process (Hallin et al. Reference Hallin, Jones, Schloter and Philippot2009; Petersen et al. Reference Petersen, Blazewicz, Firestone, Herman, Turetsky and Waldrop2012; Guo et al. Reference Guo, Deng, Qiao, Yao and Zhu2013). Correlation analysis revealed that Cu additions exhibited a clear and negative correlation with the abundance of nirK and nirS genes; such detrimental effects of Cu additions on nirK- and nirS-harbouring denitrifiers have also been reported by Magalhães et al. (Reference Magalhães, Machado, Matos and Bordalo2011). However, in our study, the overall abundance of the nosZ gene was generally not correlated to Cu amendments; this finding was contradictory to the study of Magalhães et al. (Reference Magalhães, Machado, Matos and Bordalo2011), reporting that the nosZ-bearing denitrifiers were highly sensitive to Cu, even at low concentrations. Furthermore, a possible toxicity of Cu to the entire bacterial community was found at 160μg Cu g–1 soil, as indicated by the significantly decreased copies of the bacterial 16S rRNA gene.
Taken together, our investigation revealed that soil denitrification functions and denitrifying community abundance behaved differently in sensitivity to Cu additions in different studies. This could be explained by the innate differentiations in denitrifier community abundance, diversity and composition between soils due to environmental heterogeneity (Giller et al. Reference Giller, Witter and Mcgrath1998; Bru et al. Reference Bru, Ramette, Saby, Dequiedt, Ranjard, Jolivet, Arrouays and Philippot2011; Guo et al. Reference Guo, Deng, Qiao, Yao and Zhu2013). Moreover, it could be attributed to the differences in Cu availability resulting from various soil physico-chemical properties, as total metal content, pH, texture, cation exchange capacity, minerals, organic matter and redox potential can influence the sorption, mobility and speciation of metal in soils (Giller et al. Reference Giller, Witter and Mcgrath1998; Rooney et al. Reference Rooney, Zhao and McGrath2006; Li et al. Reference Li, Sun, Huang, Ma and Zhu2010; Vega et al. Reference Vega, Andrade and Covelo2010). Although the present study was limited to only one soil, the observed inhibition of the lowest Cu addition on denitrification functions underlines the resultant sensitivity and significance of the Cu to soil denitrification process and correlated N2O emission.
3.2. Environmental implications
It is important to note that the native Cu concentration in this long-term undisturbed soil (7.9μg Cu g–1 soil) was comparable to the local background soil value (9.7μg Cu g–1 soil) of Xiamen City (China Environmental Monitoring Station 1990). In a recent study, Luo et al. (Reference Luo, Xue, Wang, Cang, Xu and Ding2015) reported an average Cu content of around 25μg Cu g–1 in soils collected from six urban parks without specific point-sources of heavy metals in Xiamen City. Thereby, in our study, it was reasonable to add Cu as low as 10μg Cu g–1 soil to simulate the increased Cu levels caused by continuous urbanisation. More broadly, the concentrations of Cu we used mainly ranged from 10μg Cu g–1 soil to 80μg Cu g–1 soil, corresponding well to the extent of urban soil Cu concentrations in Chinese cities based on their local background levels (China Environmental Monitoring Station 1990; Luo et al. Reference Luo, Yu, Zhu and Li2012; Cheng et al. Reference Cheng, Li, Zhao, Li, Peng, Qin and Cheng2014). On the other hand, urban soils could receive large N inputs, mainly in the form of NO3– rather than NH4+, and have been recognised as a potential hotspot for denitrification and N2O emission (Kaye et al. Reference Kaye, Groffman, Grimm, Baker and Pouyat2006; Li et al. Reference Li, Deng, Rensing and Zhu2014). In combination with the fact that at the national scale Cu was found to be the most significant and widespread pollutant metal in Chinese urban soils (Wei & Yang Reference Wei and Yang2010; Luo et al. Reference Luo, Yu, Zhu and Li2012), increased Cu may thus contribute to biasing urban soil denitrifier communities and profoundly alter urban soil denitrification functions.
4. Conclusions
To summarise, this study showed that Cu additions similar to increasing levels owing to rapid urbanisation can suppress soil denitrification function reflected by N2O emission and potential denitrification activity, as well as minimise the abundance of the dominant denitrifying functional genes in the investigated urban park soil. Likewise, we suggest that a small amount of Cu increase caused by urbanisation, yet bellow the grade II limit of the Chinese Soil Quality Standards (300mg Cu kg–1; State Environmental Protection Administration of China, 2008), might be favourable for mitigating emissions of the greenhouse gas N2O in urban soils. Nevertheless, how the soil denitrification process and its N2O emission patterns respond to Cu amendments may differ in different urban soils due to discrepancies in the denitrifier community as well as variation of Cu availability. In addition, conclusions obtained from incubation experiments under ideally anoxic conditions are limited to reflect real natural environments. These warrant systematic investigations as well as in situ studies. Altogether, our study offers a valuable perspective for elucidating the role that the urbanisation process plays in altering urban N biogeochemical cycling.
5. Acknowledgements
This work was financially supported by the Strategic Priority Research Program of the Chinese Academy of Sciences (XDB15020302, XDB15020402), the National Natural Science Foundation of China (41771285) and the International Science and Technology Cooperation Program of China (2011DFB91710). We thank Lars Molstad from the Norwegian University of Life Sciences for designing and programming of our robotised incubation system for analysing gas kinetics. We sincerely thank Valerie Gibson for checking the grammar of the manuscript.
