Nitrogen Uptake by Plants May Alleviate N Deposition-induced Increase in Soil N2O Emissions in Subtropical Chinese Fir Plantations

Continuous increasing nitrogen (N) deposition interferes with soil nitrogen cycle of forests, which highly impacts soil N 2 O emissions and accelerates global warming. Chinese r (Cunninghamia lanceolata (Lamb.) Hook) is one of the most widely planted species in southern China which locates in the high N deposition area. However, the impact of N deposition on soil N 2 O emissions in subtropical Chinese r plantations and the potential risk of increasing N deposition still remain elusive. Here, we conducted an in situ study in a subtropical Chinese r plantation at Fengyang Mountain Nature Reserve, China, from 2019-2020 with four different levels of N enrichment: control (CK: ambient N deposition), low-N (LN: 50 kg N ha −1 yr −1 ), medium-N (MN: 100 kg N ha −1 yr −1 ), and high-N (HN: 200 kg N ha −1 yr −1 ). We found that soil N 2 O emission rates increased with N enrichment from an average of 5.89 ± 3.66 to 20.11 ± 3.44 μg N m −2 h −1 . The N enrichment in general showed no signicant effect on the abundance of nitrate-reducing bacteria, but it tended to raise the abundance of ammonia oxidizing archaea and bacteria, and to decrease the abundance of 2 bacteria, which likely provided the microbial basis for accelerating N 2 emissions along with increasing N deposition. However, the relationship of soil N 2 emissions with input did not match an exponential but it matched a logarithmic illustrating that It found that and tended to increase leaf N concentrations and CO while that in the Chinese r plantations, and highlight that plants need to be incorporated as an important explanatory variable when predicting GHG uxes in the background of global increasing N deposition. and NO 3- -N) and biotic factors (MBC and MBN) to examine their relationships. One-way ANOVA analysis was performed to determine the dependence of pH, MBC, MBN, fresh leaf C and N, and microbial functional genes associated with soil N 2 O emissions on N addition. We used post-hoc tests (Tukey’s HSD) to examine differences among treatment levels with signicant results. All statistical analyses were performed in R 4.0.3 (R Core Team, 2020).


Introduction
Global climate change, driven by the steady increase in atmospheric greenhouse gases (GHGs) emissions, is one of the main threats to the sustainability and stability of ecosystems (L Deng et al., 2020). By 2018, the two principal GHGs nitrous oxide (N 2 O) and carbon dioxide (CO 2 ) increased by 23% and 47%, and reached about 331 ppb and 407 ppm, respectively, as compared to their levels at pre-industrial time (WMO, 2020). Forest ecosystems contributed 15-55% of the global N 2 O emissions with its area accounting for about 30.7% of Earth's total land area (U Nations, 2018). The average emission rate of N 2 O from forests to the atomosphere is 3.62±0.16 Tg N year -1 , of which 83.9% are from tropical regions (H Q Tian et al., 2016;K R Zhang et al., 2019). Therefore, the mitigation of GHG emissions from forest ecosystems is essential for balancing the GHG concentration in the atmosphere and alleviating global warming.
Aside from the direct application of nitrogen (N) used in intensive agricultural fertilization, atmospheric N deposition is the primary source of forest soil N. N deposition has been increasing as a result of human activities, such as the combustion of fossil fuels and land conversion (S-U Lena and d V Wim, 2018).
These activities have accelerated terrestrial N cycling at both the regional and global scales (D N Liu et al., 2019). N deposition has increased the concentrations of soil ammonium-N (NH 4 + -N), nitrate-N (NO 3 --N), plant N, and the leaching of inorganic N (M Lu et al., 2011). These changes directly affect nitri cation and denitri cation, and thus N 2 O emissions. Previous studies have reported that N enrichment increased (E A Davidson et al., 2004;B Koehler et al., 2009), decreased (P A Steudler et al., 2002;P Tian et al., 2018) or had no effect (S J Hall and P A Matson, 2003;A K Müller et al., 2015) on soil N 2 O emissions.
The discrepancy among these results might be due to the different responses of plant and soil microbes to changes in N availability under different nutrient conditions. For example, E A Davidson et al. (2004) reported that N 2 O emissions were weak in N-limited tropical secondary forest, due to the trees taking up extra N. S J Hall and P A Matson (1999) found that N-saturated tropical forests release more nitrogen oxides than N-limited tropical forests due to the suppressed microbial N immobilization. Similarly, N enrichment could either affect CO 2 emissions positively by increasing plant root biomass and microbial activity (L Zhou et al., 2014) or negatively by changing the decomposition of organic matter and energetic costs of N assimilation (J M Craine et al., 2007;I A Janssens et al., 2010 In situ experimental design The local ambient N deposition is estimated to be 34 kg N ha −1 yr −1 . To simulate future climate change scenarios, we set up four different levels of N enrichment (with four replicates for each treatment): a control (CK; ambient N deposition), low-N (LN; 50 kg N ha −1 yr −1 ), medium-N (MN; 100 kg N ha −1 yr −1 ), and high-N (HN; 200 kg N ha −1 yr −1 ). Each level contained four replications that were randomly assigned to eld plots. Sixteen 10 m × 10 m plots were established with adjacent plots separated by a 10 m wide buffer strip. Nitrogen was added by spraying urea (CO(NH 2 ) 2 ) solution to the forest oor starting in April 2019 for N treatments. Meanwhile, control plots received the same amount of deionized water when N was applied. The N fertilizer was applied every month with an equal split throughout the year.

Measurement of soil N 2 O and CO 2 emissions
The uxes of N 2 O and CO 2 were measured using a closed opaque static chamber (diameter ×height = 80 cm×21 cm), which consisted of a round base collar and a removable top. The top sampling chamber was covered with tinfoil to minimize temperature changes within the chamber headspace during sampling (X Zheng et al., 2020). The round base was permanently inserted 7 cm deep into the soil in April 2019. The removable top and the round base were used to form a sealed chamber headspace which is sealed with water during gas sampling to ensure air tightness. Four gas samples were taken from the chamber at 0, 10, 20 and 30 min after the bases were covered with the chambers using a 60 mL plastic syringe, the contents of which were transferred immediately into a pre-evacuated 100 mL aluminum foil gas bag. N 2 O and CO 2 concentrations of the gas samples were analyzed by gas chromatography (Agilent 7890A, Santa Clara, CA, USA) within 48 h after sampling. The uxes from the soil were calculated using the equation described by J Zhang et al. (2021a), which was based on the linear regression slope of the gas concentration during the time of chamber closure. Soil N 2 O and CO 2 emissions were sampled between 10:00-12:00 am (GMT + 8) on each sampling day. Air and soil temperatures were monitored while the gas samples were collected. The sampling frequency was twice per month. N 2 O and CO 2 uxes were sampled semimonthly from mid-May 2020 to early November 2020 (sampling was started after 13 months of N enrichment), which were equivalent to measurements at about 10 days and 24 days after N enrichment during the application of fertilizer.

Soil sampling and analysis
Topsoil (0-20 cm soil) samples were collected along with every gas sample during the study, and the soil collected from the last gas sample was used for the microbial functional gene analysis (bagged on dry ice and stored at −80 °C). Four soil cores were randomly collected from each plot and mixed into one composite sample, and then frozen before being passed through a 2 mm sieve. Soil NH4+-N and NO  Vance et al., 1987). We measured the contents of MBC and MBN in soil at the middle (early-August) and end (early-November) of the growing season. Gravimetric soil water content was determined by drying in an oven at 105 °C until the soil reached a constant weight. The pH of soil samples collected in November 2020 were measured at a soil-to-water ratio of 1:2.5 (v/v). We also collected fresh leaves in November 2020, and we determined the concentrations of total C and total N with an elemental autoanalyzer (Vario MAX CN, Elementar Analysensysteme, GmbH, Langenselbold, Germany).

Quanti cation of microbial functional genes
Total DNA was extracted from 0.5 g soil with FastDNA SPIN Kit for soils (BIO 178 101, Qbiogene, Carlsbad, CA, USA), following the manufacturer's protocol. We used 0.8% agarose gel electrophoresis to detect the quality and integrity of the extracted DNA. The quanti cation of functional marker genes AOA, AOB, nirK, nirS, and nosZ was performed by real-time quantitative PCR using a CFX96 Optical Real-Time Detection System (Bio-Rad Laboratories Inc., Hercules, CA, USA). Information regarding the gene-speci c primers and thermal conditions can be found in our previous study X Zheng et al. (2020). Each reaction mixture (25 μl) consisted of 12.5 μl 1 × SYBR Premix Ex Taq (Takara, Tokyo, Japan), 0.25 μl of each primer (10 mM), and 1 μl of template DNA (1-10 ng). Standard curves were prepared using a serial dilution of known copy number plasmid DNA from one representative clone containing each target gene.

Statistical analyses
We used mixed model analysis of variance (ANOVA) to examine the dependence of soil N 2 O and CO 2 emission rates, soil temperature, soil moisture, NH4+-N, and NO 3 --N on N addition with time as a random effect. We performed pairwise correlation analyses between soil N 2 O and CO 2 emissions among soil abiotic factors (soil temperature and moisture, and NH4+-N and NO 3 --N) and biotic factors (MBC and MBN) to examine their relationships. One-way ANOVA analysis was performed to determine the dependence of pH, MBC, MBN, fresh leaf C and N, and microbial functional genes associated with soil N 2 O emissions on N addition. We used post-hoc tests (Tukey's HSD) to examine differences among treatment levels with signi cant results. All statistical analyses were performed in R 4.0.3 (R Core Team, 2020).

Effects of N enrichment on soil characteristics and N 2 O and CO 2 emissions
During the experimental period from May to November, the mean air temperature (7.48 to 23.32 °C) and soil temperature (9.75 to 20.07 °C) was 18.08 °C and 15.83°C, respectively (Fig. 1a, 1b). N enrichment signi cantly increased soil NH 4 + and NO 3 concentrations (Figs. 1d, 2a and 2b), but it decreased soil moisture (Figs. 1c and 2c) and soil pH (Fig. 2d). For example, soil moisture decreased from 85% in the controls to 70.9% in the high-N plots (Fig. 2c). N 2 O emission rates increased logarithmically with N enrichment from an average of 5.89 ± 3.66 to 20.11 ± 3.44 μg N m −2 h −1 (Fig. 3a). Only high-N enrichment signi cantly increased soil N 2 O emission rates ,compare with control (Fig. 3a). CO 2 ux showed increasing trends with elevated N enrichment from 122.15 ± 14.86 to 218.29 ± 32.43 mg N m −2 h −1 , but the effect of N enrichment was not statistically signi cant (Fig. 3b).

N effects on microbial biomass and functional genes associated with soil N 2 O emissions
Over the study period, N enrichment signi cantly decreased soil MBC and MBN in the mid-growing season but it had no effect during the end of the growing season (Figs. 4a and 4c). The microbial biomass at the end of the growing season was signi cantly higher than that in the mid-growing season ( Figs. 4a and 4c). Fresh leaf N concentration tended to increase with N enrichment (Fig. 4b), while the concentration of leaf C remained consistent (Fig. 4d). The N enrichment signi cantly increased the abundance of AOA and AOB (Figs. 5a and 5b), and signi cantly decreased the abundance of nosZ ( Fig.  5e) but it had no effect on the abundance of nirS or nirK (Figs. 5c and 5d).

Relationships between environmental factors and soil N 2 O and CO 2 emissions
Among the environmental factors we measured, soil N 2 O emission rates were negatively correlated with MBC and MBN, but there were no relationships with soil temperature, soil moisture, air temperature, NH4+- . As far as our study is concerned, it remains unclear which populations of ammonia-oxidizers (AOA and AOB) are primarily responsible for nitri cation in forest soils, or which contributed more N 2 O in response to N enrichment.
Nitrogen enrichment decreased the number of nosZ gene copies (Fig. 5e) We found that soil microbial biomass was signi cantly negatively correlated with NH 4 +-N, and that soil NO 3 --N was positively correlated with MBC ( Figs. 6a and 6b). The content of NO 3 --N in the high-N and mid-N treatments was signi cantly higher than that in low-N, but there was no difference in NH 4 + -N among the three treatments. Our results suggested that nitri cation may be the main source of N 2 O emissions, which caused an accumulation of nitrate in the soil. Soil moisture is crucial for gas diffusion and plays a dominant role in controlling nitri cation and denitri cation, which in turn affects With respect to the abundance of N 2 O emissions related functional genes, the N enrichment in general showed no effect on the abundance of nitrate-reducing bacteria, but it raised the abundance of ammonia oxidizing archaea and bacteria, and decreased the abundance of N 2 O-reducing bacteria, providing the microbial basis for accelerating soil N 2 O emissions along with increasing N deposition (Fig. 5).
Therefore, we speculate that the relationship between N 2 O emission and N input may be exponential growth. However, in contrary to our speculation, we found that the soil N 2 O emissions increased logarithmically with N input (Fig. 3a), illustrating that the risk of increasing N deposition on soil N 2 O emissions was attenuated.
The IPCC assume a linear relationship between N application rate and N This may suggest that the Chinese r plantation was N-limited in our study area, which was consistent with L Li et al. (2019).
Under the N-limited system in our study, the competition of available N (NH 4 + and NO 3 -) by plants uptake may be a major reason for the slower than linear rate of soil N 2 O emissions with increasing N input. In this study, N enrichment not only stimulated plant growth, but also increased leaf N content. We also observed that N enrichment tended to increase fresh leaf N concentration, even though these increases were not signi cant (Fig. 4b). be the reason why the leaf N content did not respond to different levels of N enrichment, which was supported indirectly by the decreased soil water content as N enrichment increased. E A Davidson et al. (2004) showed that N enrichment increased tree biomass and foliar N concentration, which resulted in no clear response of N 2 O emissions to N enrichment. Taken together, the plant's absorption of available N from the soil effectively reduced the N 2 O emissions caused by N enrichment in the N-limited Chinese r plantation. However, our results are based on short-term nitrogen enrichment treatments, and thus longterm observations of the response of Chinese r plantation GHGs to N enrichment are necessary to more fully understand the long-term impacts of N enrichment.
The soil moisture decreased markedly with N enrichment in our study (Figs. 1c and 2c), which may indirectly support the competition of plants with microbes to uptake N nutrients together with soil water (B Koehler et al., 2009;A K Müller et al., 2015;X Zheng et al., 2020). It has been reported that N enrichment can increase plants transpiration while decreasing soil water leaching below the rooting zone (X Lu et al., 2018). This phenomenon may have resulted from promoted plant growth due the increased availability of N (J Jiang et al., 2019), which increased the absorption of water by roots and transpiration, and thus decreased soil moisture.
Soil microorganisms are sensitive to changes in the availability of soil N (C Wang et al., 2018). Previous studies reported that N enrichment could increase microbial biomass by relieving N limitation (Z Zhou et al., 2017) or inhibit microbial biomass by decreasing soil pH (K K Treseder, 2008). N enrichment signi cantly decreased soil pH (Fig. 2d) agreed with D S Tian and S L Niu (2015), who found that soil pH decreased by 0.26 on average globally due to N enrichment and enhanced Al 3+ concentrations in soils. The lower soil microbial biomass under increased N enrichment observed in our study (Figs. 4a and 4c) may have resulted from enhancements in the content of Al 3+ due to lower pH, and thus caused Al 3+ poisoning in microbes (W D Bowman et al., 2008). Interestingly, we found that the biomass in the middle of the growing season was signi cantly lower than that in the end of the growing season (Figs. 4a and 4c), which was inconsistent with the view that microorganisms were at the optimal temperature (J Pietikäinen et al., 2005). Indeed, under N-restricted conditions, the results of the isotopic labeling of N 15 indicated that plants can in fact successfully compete for N with microbes after N enrichment (J P Kaye and S C Hart, 1997). This nding indicated that the competition between plants and microorganisms intensi ed plant growth in the vigorous growth stage, was alleviated in the slow growth stage, and thus promoted the growth of microorganisms. In addition, soil N 2 O emission rates were negatively correlated with soil microbial biomass (MBC and MBN; Tab. 2). S J Hall and P A Matson (2003) found that microbial N immobilization was part of the reason for the low losses of soil N-oxides following N fertilization. Thus, microbial N xation can mitigate N 2 O emissions to some extent, and the mitigation effect is affected by plant and soil pH.
During our observation period, N enrichment increased soil CO 2 emissions, but the differences caused by N enrichment were not statistically signi cant (Fig. 3b).  2018), which suggested that the response of heterotrophic respiration to N enrichment may decrease or remain unchanged rather than increase. Conversely, we observed that N enrichment increased soil CO 2 emissions (Fig. 3b), which indicated that there was a dominant contribution by plant root growth to increased CO 2 emissions, as evidenced by the decreased soil moisture (Fig. 2c).

Conclusions
Elevated N enrichment increased N 2 O emissions following a logarithmic trend in N-limited Chinese r plantations. Nitrogen enrichment promoted plant N immobilization by increasing plant growth and leaf N concentration, which therefore dampened the increased N 2 O emissions caused by N enrichment. N enrichment enhanced the abundance of AOA and AOB, which may indicate that nitri cation is the main source of N-stimulated N 2 O emissions. Considering the increased deposition of N in Chinese r plantations due to human activity, our ndings are of great importance for forest management. However, long-term studies are needed to more fully explore the threshold and the cumulative effects of increased N deposition on GHG uxes in Chinese r plantations.

Declarations
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