Effect of mowing on N₂O and CH₄ fluxes emissions from the meadow-steppe grasslands of Inner Mongolia

Abstract

To assess the impacts of mowing on N₂O and CH₄ fluxes emissions from the meadow-steppe grasslands of Inner Mongolia, China, two regimes were investigated: unmown since 2005 (UM), and mown once every three years since 2009 (M3). On-site measurements were conducted continuously during a year-round period (August 2011 to August 2012). During the observation period, three diurnal cycles were also measured. In addition, a targeted laboratory experiment was conducted to make up for the few measurements in winter. A large pulse of N₂O emissions related to freeze-thaw cycles was observed at M3 during the spring thaw. Results showed that the meadow-steppes played a role as a sink for CH₄ and a source for N₂O. Significantly lower mean CH₄ uptake at UM (40.3 μg C·m^-2·h^-1) as compared to M3 (70.5 μg C ·m^-2·h^-1) (p<0.01), and significantly higher mean N₂O efflux at UM (6.3 μg N·m^-2·h^-1) as compared to M3 (4.3 μg N·m^-2·h^-1) (p<0.05) were found. The laboratory experiment results revealed that mowing changed the soil conditions that favor the activity of denitrifiers during thawing periods. The CH₄ and N₂O fluxes were significantly correlated with soil temperature (p<0.05). Mowing affected CH₄ uptake and N₂O emission mainly through its effect on vegetation types and some soil properties, such as soil inorganic N content, soil temperature, and soil moisture content, while soil inorganic N and moisture were not leading factors. Our results also suggested that mowing could mitigate the potential global warming in terms of CH₄ uptake and N₂O emissions.

Keywords

mowing / meadow-steppe / methane / nitrous oxide / freeze-thaw cycles / global warming

Introduction

Methane (CH₄) and nitrous oxide (N₂O) are the most important greenhouse gases in the atmosphere, the radiative forcing of which are, on a molar basis, about 20–30 and 220–290 times greater than that of CO₂, respectively ( Rodhe,1990; Lloyd, 1995). The atmospheric concentrations of CH₄ and N₂O have increased rapidly due to anthropogenic activities, such as land use/cover changes, agriculture, and fossil fuel use ( Ojima et al., 1993; Houghton et al., 1999; Dlugokencky et al., 2003; IPCC, 2007; Zhou et al., 2007a; Zhang et al., 2014a). Because of these human activities, the global atmospheric concentrations of CH₄ and N₂O have increased from a pre-industrial values of approximately 715 ppb and 270 ppb to current values of 1,774 ppb and 319 ppb, respectively ( IPCC, 2007). Furthermore, N₂O emissions contribute to the catalytic depletion of the ozone layer via reactions with singled oxygen atoms ( Crutzen, 1970; Ravishankara et al., 2009). Therefore, more attention should be paid to the increment of CH₄ and N₂O concentrations in the atmosphere.

Grassland covers roughly 25% of the earth’s land surfaces, and is one of the most important global terrestrial ecosystems. The natural grassland of China covers an area of 3.9 × 10⁸ ha, and the Leymus chinensis steppe, which is the dominant grassland type of the semi-arid grassland area in Inner Mongolia, covers an area of approximately 90,000 km² ( Wang, 2004). In the semi-arid grasslands of Hulunber, Inner Mongolia, grazing and mowing are important management practices. Wolf et al. (2010) carried out a year-round study with high and low temporal resolution at ten steppe grassland sites in Inner Mongolia of China, and concluded that grazing decreased rather than increased N₂O emissions. Smith et al. ( 2000) reported that the conversion of native grasslands into pastures decreased the sink strength of soils for atmospheric CH₄. These land use and management changes will directly affect CH₄ uptake and N₂O emission. Therefore, variation, even if minor, in the exchange of these trace gases (N₂O and CH₄) between grassland ecosystems and the atmosphere can be of significance for the global atmospheric budgets.

Mowing, an important human practice in the Eurasian steppe management, is increasingly being used as a method to collect forage to feed livestock, replacing traditional practices ( Foster et al., 2009). Huston ( 1979) and Collins et al. ( 1998) found that mowed plots had more than double the species richness of plots that were not mown, and mowing in anthropogenically stressed grasslands enhanced biodiversity. Meanwhile, mowing also has various other effects on grassland ecosystems, e.g., changes in soil temperature and moisture, soil microbial growth, and biomass ( Parr and Way, 1988; Wan et al., 2002; Bahn et al., 2006; Robson et al., 2007; Zhou et al., 2007b; Ilmarinen and Mikola, 2009; Luo et al., 2009; Gavrichkova et al., 2010). Soil–atmosphere exchange of CH₄ and N₂O relies on a suite of complex processes: production, consumption, and diffusion processes in the sequential biochemical reactions. In general, the magnitude of their uptake or emission depends on soil physical properties, environmental conditions, and soil biologic characteristics such as soil temperature and moisture content, soil mineral N content, microbial activity and population size, and vegetation type ( Willison et al., 1997; Smith et al., 2000; Groffman et al., 2006; Maljanen et al., 2007; Chen et al., 2010). Some of the changes induced by mowing have been shown to influence CH₄ and N₂O fluxes in grassland ( Wang et al., 2005; Du et al., 2006; Blagodatsky and Smith, 2012; Cai, 2012). Thus, in the present study we hypothesized that mowing would affect CH₄ and N₂O fluxes by: (i) affecting the soil inorganic N content and species composition, (ii) changing the soil moisture content, and (iii) the soil temperature.

Only a few studies have been conducted on the effect of mowing on greenhouse gas fluxes from the meadow-steppe grasslands of Inner Mongolia, and the effects on CH₄ and N₂O fluxes in particular are not well understood ( Guo et al., 2010, 2011a, b; Zhang et al., 2014b; Zhong et al., 2014). In these previous studies, however, the measurements of grassland greenhouse gas fluxes were often performed over short time periods, and mostly during the growing season (from May to September). Thus, the estimations of emissions feature high levels of uncertainty due to the absence of measurements during the non-growing seasons in which the seasonal and repeated freeze-thaw events occur. Freeze-thaw cycles affect C and N dynamics in soils, and thus also have a considerable impact on the emissions of GHGs ( Zhu et al., 2009). Mosier et al. ( 1996) reported that winter fluxes contributed 20%–40% of the annual N₂O emissions and 15%–30% of the CH₄ consumption in North American prairie systems. Röver et al. (1998) found that N₂O-emission-related freeze-thaw cycles could contribute up to 70% of the annual N₂O emissions. In our field measurements, CH₄ uptake peaks at both sites were not found, and a large pulse of N₂O emissions was observed at mown sites, but not at unmown sites, during the spring thaw. To better understand this phenomenon of the freeze-thaw effect on the soil N₂O flux at mown meadow-steppe sites, a laboratory incubation experiment under controlled conditions was carried out.

The hope in conducting these field and laboratory measurements of CH₄ and N₂O fluxes in response to mowing on the meadow-steppe grassland of the Hulunber high plain in Inner Mongolia was that we can contribute to an improved understanding and more accurate evaluation of the global carbon budget and climate change. In this context, the aim of the study was threefold: (1) to study the effect of mowing on environmental factors; (2) to examine the effect of mowing on CH₄ and N₂O fluxes from the meadow-steppe grassland; and (3) to calculate the net effect on potential global warming in terms of net CH₄ and N₂O emissions.

Materials and methods

Site descriptions and field treatments

The experimental site was located in a meadow-steppe grassland adjacent to the Hulunber Grassland Ecosystem Research Station of the Chinese Academy of Agricultural Sciences in Inner Mongolia, China (49°19′N, 120°03′E; 628 m a.s.l.). The climate in this region is semi-arid, with a frost-free growing season of 95–110 days from early May to early September. Annual mean precipitation is 400 mm (150–550 mm), with a large inter-annual variation, falling primarily between June and August, which coincides with the growing season. Annual mean air temperature is -2°C to 1°C, with a minimum monthly mean of -26°C in January, and a maximum of 21°C in July. The region has a long winter period of nearly six months, from October to April of the following year, a short autumn (September; Julian days 241–270), spring (May, June; Julian days 120–180), and summer (July, August; Julian days 181–240) ( Du et al., 2006). The soil in this region is frequently exposed to the episodes of intensive freeze-thaw cycles, starting to freeze in late October, and thawing in April. The unmown steppes (UM) have been fenced since 2005, while for the mown steppes a machine has been used since 2009 to mow the grasses in late August once every three years (M3). Grasses are mowed at a height of 5 cm, and a control of unmown grass (height of approximately 25 cm) is maintained. The study site’s soil is chernozem with an average thickness of 30–40 cm. The major soil characteristics are provided in Table 1. The meadow-steppe grassland is dominated by Leymus Chinensis.

In situ measurements of CH₄ and N₂O fluxes

The CH₄ and N₂O flux measurements were carried out from August 2011 to August 2012. Four replicate plots were simultaneously observed for each field treatment. Gas samples were collected using the static chamber method. The static chamber was made of stainless steel and equipped with a fan, 10 cm in diameter, installed on the top wall of each chamber to ensure complete gas mixing. The chamber consisted of two parts: a square box (without a top and bottom; dimensions: 0.5 m × 0.5 m × 0.1 m (length × width × height)) and a box with a removable cover (without a bottom; dimensions: 0.5 m × 0.5 m × 0.5 m (length × width × height)). The square box was inserted directly into the soil at about 10 cm below the soil surface, and the cover was placed on top during sampling times and removed afterwards. A white adiabatic cover was added outside of the stainless steel cover to reduce the impact of direct radiative heating during sampling. The chamber was closed for almost an hour and five gas samples were collected every 10 min using plastic syringes fitted with three-way stop-cocks. The collected gas samples were analyzed with a gas chromatograph (Agilent 7890 GC USA) within 24 hours. The gas chromatograph was equipped with a flame ionization detector (FID) for CH₄ analysis, and an electron capture detector (ECD) for N₂O analysis. The gas chromatography configurations for analyzing concentrations of CH₄ and N₂O, and methods for calculating the fluxes of CH₄ uptake and N₂O emission, were the same as those reported by Wang and Wang ( 2003). The flux measurements were taken once or twice per week during the growing season, and once or twice per month during the non-growing season. Intensive flux measurements at two-hour intervals to investigate the diurnal variations of the meadow-steppe were taken on 10 August 2011, 18 June 2012, and 10 August 2012, respectively. Annual GHG emission or uptake was estimated by calculating cumulative fluxes over an experimental period. Based on the equation of global warming potential (GWP), an estimate of the degree of contribution to global warming, the CO₂ equivalent [CO_2(e)] of the net emission of GHGs (only CH₄ and N₂O were taken into account) was determined using a transformation coefficient of 298 for the N₂O emission flux and 26 for the CH₄ uptake flux ( Li et al., 2004; IPCC, 2007). The CO_2(e) was calculated using the following equation:

\mathrm{C}{\mathrm{O}}_{\mathrm{2}\left(\mathrm{e}\right)}={\mathrm{N}}_{\mathrm{2}}\mathrm{O}÷\mathrm{28}\times \mathrm{12}\times \mathrm{298}+\mathrm{C}{\mathrm{H}}_{\mathrm{4}}÷\mathrm{44}\times \mathrm{16}\times \mathrm{26},

where CO_2(e) (kg C·ha^-1·yr^-1) is the CO₂ equivalent of net emissions of GHGs, and N₂O and CH₄ are the N₂O flux (kg N·ha^-1·yr^-1) and the CH₄ flux (kg C·ha^-1·yr^-1), respectively.

Laboratory experiment

The incubation experiment was performed in a 360 mL flask. Tap water (directly extracted local groundwater) was used to just flood the top of the soil to simulate field conditions. The flask still had sufficient headspace left at the top of the intact soil for gas sampling. Each flask was then sealed with an airtight lid and placed in a freezer to simulate freeze-thaw cycles. To begin with, the soil was frozen for 1 week at -15°C, and then we raised the temperature up to 5°C for 1 week. In subsequent cycles, the temperature was changed from -10°C to 5°C, from -5°C to 10°C, and from -5°C to 5°C. After a temperature change, the soil cores were allowed to equilibrate to the new temperature at least 12 hours before the start of flux measurements.

Gas samples from the headspace of the flasks were taken with 60 mL plastic syringes fitted with three-way stopcocks at 12 hours, 24 hours, 3 days, and 7 days. The gas samples were immediately analyzed using gas chromatography (Agilent 7890 GC USA). The gas chromatography was equipped with an electron capture detector for N₂O analysis. N₂O fluxes were calculated using equation ( Du et al., 2006)

f={\displaystyle \frac{\mathrm{\Delta }m}{M\mathrm{\Delta }t}}={\displaystyle \frac{\rho v\mathrm{\Delta }c}{M\mathrm{\Delta }t}},

where ρ (ppm) is the density of the gas, ∆c/∆t is the variation ration of the gas concentration, M (g) is the mass of the soil, and v (mL) is the volume of the headspace in the flask.

Measurement of environmental factors

During each gas sampling process, the internal temperature of the chamber and the soil surface temperature were simultaneously recorded, using a portable digital thermometer (JM624, Tianjin, China). Soil temperature and moisture at depths of 10 and 20 cm were continuously monitored by an automated measuring system using a Hobo Micro Station Data Logger (H21-002, USA) every 10 seconds, and were stored every 1 minute during the growing season, and every 1 minute and every 5 minutes during the non-growing season. Climatic data (daily precipitation, air temperature) were obtained from a local meteorological station.

Soil sampling and vegetation description

Soil cores (0–20 cm layer) were collected nine times using a steel soil corer (7 cm in diameter): in spring (15 May, 31 May, 15 June, 30 June), summer (13 July, 1 August, 16 August, 27 August), and in autumn (20 September). Five soil cores were randomly taken for each site at each time. Intact soil was separated into five depth intervals (0–3 cm, 3–6 cm, 6–9 cm, 9–12 cm, 12–15 cm) for laboratory incubation in a 360 mL flask. For soil analysis, each soil sample was separated into three depth intervals (0–5 cm, 5–10 cm, 10–20 cm), and then five soil samples were mixed evenly by interval for determination of soil organic matter (SOM), soil \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N concentrations. Soil \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N concentrations were determined photometrically after extracting the soil with 1 mol·L^-1 KCl solution (1:2 w:w) for 1 h. Organic matter content was determined by the soil bath-K₂CrO₇ titration method.

The grassland vegetation was observed and measured once a month (from May to September) using six quadrats of 1 m² in each site. Plant standing biomass was harvested, oven-dried at 65°C for 48 hours, and weighted for each species ( Zhong et al., 2014).

Statistical analysis

The statistical software packages SPSS 17.0 (SPSS Inc., Chicago, USA) and Origin 8.0 (Origin Laboratory Corporation, USA) were used. Differences in CH₄ and N₂O fluxes, aboveground biomass, average soil temperature, soil moisture, SOM, soil \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N among treatments were determined using ANOVA. Pearson correlation was used to determine correlations between CH₄ and N₂O fluxes and the environmental factors. All statistical tests were performed at a significance level of level of α = 0.05.

Results

Environmental conditions and soil conditions

The daily means of air temperature, precipitation, and soil temperature and moisture at 10 and 20 cm depth at the two sites are given in Fig. 1. The average air temperature was -1.7°C, ranging from -37.8°C (winter) to 25.5°C (summer). Annual precipitation (247.6 mm) was lower than the long-term average (400 mm), and seasonally precipitation in spring and summer was higher. The relatively high and low daily precipitation generated comparatively wet and dry conditions, which altered soil moisture and soil temperature. In winter, heavy snow generated soils with saturated conditions, and the freeze period was from roughly the first 5 days of October (Julian days 305) to the middle 15 days of April the following year (Julian days 105) (Fig. 1(a)). The surface snow and ice then began to thaw with great fluctuation, and totally thawed in the first 5 days of May (Julian days 120) as temperature increased.

Mowing affected soil temperature at 10 cm depth in summer and winter, soil temperature at 20 cm depth in winter significantly (p<0.01), but no significant effects were found on soil temperature in other seasons. However, there was a significant effect on soil moisture in all seasons except at 20 cm in autumn (Table 2).

Soil analysis and vegetation conditions

Both \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N, \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N and SOM showed the same distribution, decreasing with increasing depth, i.e., the \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N of soil at depths of 0–5, 5–10 and 10–20 cm was 5.07, 4.46, and 4.44 mg·kg^-1, respectively (Table 1). There was no significant difference between M3 and UM in terms of \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N, \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N, and SOM, indicating that inorganic nitrogen ( \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N) and SOM restored gradually after mowing, but their concentrations were higher in UM than M3 fields.

Aboveground biomass showed strong seasonal fluctuations, with peak values in July (Fig. 2). Mowing had no effect on aboveground biomass (p>0.05); however, a marginally significant effect of mowing on species composition was found (p<0.01) (data not shown). The mown meadow-grassland was found to be dominated by Leymus Chinensis, whereas the dominant species of unmown grassland-meadow was Carex tristachya, indicating moderate mowing can prevent grassland degradation in meadow-steppe grassland.

Diurnal variation in soil CH₄ and N₂O fluxes

The diurnal cycle in three different growth stages of grass was analyzed and was found to vary greatly with no stable pattern (Fig. 3). During the panicle initiation stage, diurnal variation of CH₄ uptake was slight. The daily average CH₄ uptake flux was 63.1 μg C·m^-2·h^-1 in UM, which was rather lower than in the M3 steppe, and mean uptake flux was 75.1 μg C·m^-2·h^-1, reaching a maximum at 16:00 and falling to a minimum at 20:00. Figures 3(a) and 3(c) suggest that there was a great variation in CH₄ uptake, even when observed in the same growing stage. Figure 3 also shows that the timing of the peak uptake varied in different stages, but on a diurnal flux basis, average CH₄ uptake over the time period of 08:00–14:00 would be representative for a diurnal period.

The average values of CH₄ uptakes during the whole day, daytime and nighttime for UM and M3, are shown in Table 3. The maximum CH₄ uptake was observed with a daily average of up to 63.1 μg C·m^-2·h^-1 for UM and up to 87.5 μg C·m^-2·h^-1 for M3. In general, the daily mean CH₄ uptake flux from the M3 meadow-steppe tended to be roughly 1.58-fold greater than that of the UM field. Meanwhile, mowing changed the proportion of CH₄ flux during daytime and nighttime, i.e., 64% and 36% for UM and 54% and 46% for M3 on 7 and 8 August, respectively. As shown in Table 4, mowing enhanced CH₄ uptake mainly through increasing its flux during nighttime, e.g., in the panicle initiation stage, the average CH₄ uptake flux at night was 54.1 μg C·m^-2·h^-1 for UM and 72.6 μg C·m^-2·h^-1 for M3.

Similar to the CH₄ uptake, no stable pattern of diurnal variation in N₂O emission by the UM and M3 steppes was revealed, with mean emissions of 3.7 and 0.5 μg N·m^-2·h^-1 respectively in the panicle initiation stage (Fig. 4, Table 3). Figures 4(a) and 4(c) suggest that there was a great inter-annual variation in N₂O uptake (3.5 μg N·m^-2·h^-1, and 7.6 μg N·m^-2·h^-1 for UM in 2011 and 2012, respectively). These variations all implied that any low-frequency diurnal observation may greatly under- or over-estimate the total N₂O emission, due to the enormous variability.

On a diurnal flux basis, the average N₂O emission over the time period of 08:00–10:00 would be representative of a diurnal period. The data from diurnal-cycle measurements conducted in various growth stages (Table 3) indicated that the daily average N₂O emission fluxes from the UM and M3 steppes varied greatly from 0.5 to 7.6 μg N·m^-2·h^-1, and the daily mean N₂O emission by the UM steppe tended to be roughly 2.7-fold greater than that of the M3 field (means of 4.9 μg N ·m^-2·h^-1 and 1.9 μg N·m^-2·h^-1, respectively). In contrast to the influence of mowing on CH₄, mowing decreased N₂O emissions by changing the proportion of its flux during daytime and nighttime, i.e., 55% and 45% for UM, and 60% and 40% for M3 on 18 and 19 June, respectively.

Seasonal variation in soil CH₄ and N₂O fluxes

Figure 5 shows that the year-round measured CH₄ uptake fluxes were usually negative. This suggests that the meadow-steppe of the Inner Mongolia played a role as a sink for the atmospheric CH₄. At the same time, the CH₄ uptake of either site showed a large seasonal variability (Fig. 6). The maximum CH₄ uptake was observed in summer or autumn, with means reaching -102.2 μg C·m^-2·h^-1 for UM and -220.7 μg C·m^-2·h^-1 for M3. During winter, CH₄ uptake at the two sites was still observed with a range of -2.23-42.8 μg C·m^-2·h^-1 (Fig. 5). As shown in Table 5, the CH₄ uptake at the M3 steppe was 1.75 times higher than that at the UM site, with the annual average reaching 70.5 and 40.3 μg C·m^-2·h^-1, respectively. The CH₄ uptake fluxes at M3 averaged -52.8 μg C·m^-2·h^-1, -116.2 μg C·m^-2·h^-1, -118.9 μg C·m^-2·h^-1, and -55.6 μg C·m^-2·h^-1 during the seasons of spring, summer, autumn and winter, respectively. At UM, the average flux during each season was -33.0 μg C·m^-2·h^-1, -54.9 μg C·m^-2·h^-1, -44.6 μg C·m^-2·h^-1, and -37.6 μg C·m^-2·h^-1, respectively. Accumulatively, the former was higher than the latter by 4% in spring, 14% in summer, 9% in autumn, and 15% in winter. This indicated that mowing drastically enhanced CH₄ uptake in summer, autumn and winter.

As Fig. 5 shows, the year-round measured N₂O emission fluxes from the M3 and UM meadow-steppes were usually positive, with some sporadic N₂O uptake events, especially during soil freeze-thaw periods. This suggests that the meadow-steppe of the Inner Mongolia played a role as a source of the atmospheric N₂O, with an annual emission flux of 4.2 and 6.2 μg N·m^-2·h^-1 for the M3 and UM steppe, respectively (Table 4). There was also an obvious difference in N₂O emission intensity between different seasons (Fig. 6). The N₂O emission at M3 averaged 2.2 μg N·m^-2·h^-1, 3.9 μg N·m^-2·h^-1, 11.0 μg N·m^-2·h^-1, and 4.0 μg N·m^-2·h^-1 during the seasons of spring, summer, autumn, and winter, respectively. On average, the N₂O emission at M3 was lower than that at UM by 2.4 μg N·m^-2·h^-1 in spring, 2.6 μg N·m^-2·h^-1 in summer, 6.2 μg N·m^-2·h^-1 in autumn, and 1.1 μg N·m^-2·h^-1 in winter. Accumulatively, the latter were 2.1, 1.7, 1.6, and 1.3 times higher than the former in spring, summer, autumn, and winter, respectively. This indicates that mowing decreased N₂O emissions by mainly having an effect in spring, summer, and autumn.

Laboratory experiment

Measurements of N₂O in the gas profile showed that the soil N₂O concentration never reached zero and was generally small in frozen soils, and the highest soil N₂O concentration was generally observed during soil thawing periods (Fig. 7). In UM, N₂O emissions increased (from 0.001 μg·g^-1·h^-1 to 0.03 μg·g^-1·h^-1) at the start of soil thawing only, and declined strongly and remained constant during the subsequent continuation of thawing. In M3, the N₂O emission was characterized by a fast increase from 0.002 μg·g^-1·h^-1 to 0.05 μg·g^-1·h^-1 followed by a rapid decline to 0.02 μg·g^-1·h^-1 2 days later from -15°C to 5°C, and an increase from 0.004 μg·g^-1·h^-1 to 0.04 μg·g^-1·h^-1 followed by a rapid decline to 0.02 μg·g^-1·h^-1 3 days later from -10°C to 5°C during the thawing period, before finally remaining constant. The spring-thaw N₂O flux in M3 was larger than in UM, which was in agreement with our field findings. Measurements of N₂O in the gas profile also revealed that the observed peak N₂O emission during spring thaw occurred in the topmost layer (0–6 cm).

The soil \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} concentration decreased during the experiment (Table 4). The nitrate concentration after the freeze-thaw cycle experiment was 6.90 mg·kg^-1 in UM and 9.14 mg·kg^-1 in M3 (8.32 mg·kg^-1 and 11.15 mg·kg^-1 at the beginning, respectively). The soil \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} concentration increased from 4.45 mg·kg^-1 to 12.12 mg·kg^-1 in UM, and from 4.34 mg·kg^-1 to 12.86 mg·kg^-1 in M3 during the whole experiment. The total N concentration and SOM showed little change.

Net effect on potential global warming in terms of net N₂O and CH₄ emissions

The annual net effect of CH₄ uptake and N₂O emission was determined by the sum of CO_2(e), which was calculated by multiplying the global warming potential with the annual total net emissions of N₂O and CH₄ (i.e., 298 and 26 for the unit mass of N₂O and CH_4, respectively, at the 100-yr time scale ( IPCC, 2007)). The calculated results listed in Table 5 showed that the CO_2(e) of the annual net emissions of CH₄ and N₂O from the UM and the M3 steppe sites was 36.1 and -10.3 kg C·ha^-1·yr^-1, respectively. This indicates that the mowing of the meadow-steppe mitigated the potential global warming; however, the role of mowing on the meadow-steppe in the terms of the atmospheric greenhouse effect has rarely been considered.

Discussion

N₂O emissions from frozen soil and during thawing

In winter months, the grasslands in this study are usually subject to heavy snowfall and experience several days or weeks of subzero temperature. The grasslands are exposed to freeze-thaw cycles for a period of time. Müller et al. ( 2002) reported that N₂O emissions occurred even when the soil was frozen. Our results also demonstrated that N₂O emissions could be measured for several days at both sites during the period of soil freezing, indicating that soil microorganisms were still active in frozen soils. The N₂O emission peaks were observed after soil thawing, which lasted only for a few days, and the emission of N₂O dropped off very quickly at the onset of the freezing cycle in M3 (Figs. 5 and 7). Meanwhile, the emissions decreased in the subsequent thawing periods (Fig. 7). There was no matching peak of N₂O emissions at the UM site in the field measurements. Contrary to expectations, the accumulation of N₂O in winter accounted for 70% of the annual accumulation for the UM site and 55% of the annual accumulation for the M3 site (Table 5). The results do not contradict our laboratory findings, as in our field measurements N₂O emissions were lower at M3 than UM in frozen periods, and only higher in a short-lived thawing periods. Therefore, mowing had changed the soil and environmental conditions that determine emissions of N₂O during thawing periods. The \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} concentration increased by approximately 7.67 mg·kg^-1 at UM and by 8.52 mg·kg^-1 at M3. Meanwhile, the \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} concentration decreased by approximately 1.42 mg·kg^-1 at UM and 2.01 mg·kg^-1 at M3, respectively (Table 4). N₂O emissions peaks occurred as a consequence of favorable denitrification conditions due to the death and lysis of microbial cells and/or the destruction of soil aggregates in saturated soil as a result of snowmelt ( Groffman and Tiedje, 1989; Christensen and Christensen, 1991; Herrmann and Witter, 2002; Müller et al., 2002).

Effects of mowing on soil CH₄ and N₂O fluxes

Mowing reduced soil moisture, which acted as a diffusion barrier for methanotrophic microorganisms and determined the proportion of anaerobic/aerobic conditions in the soil. As Table 2 shows, there was higher soil moisture at UM (mean of 9.5%, v/v) than at M3 (mean of 6%, v/v), especially in spring and summer (p<0.01). Although positive correlations between CH₄ uptake and soil moisture have also been reported in several other studies ( Smith et al., 2000; Jones et al., 2005; Livesley et al., 2009), our results showed no significant correlation between soil moisture and CH₄ uptake during the year-round measurements (p>0.05). This suggested mowing enhanced CH₄ uptake (70.5 versus 40.3 μg C·m^-2·h^-1) by affecting other more important factors, not primarily through the effect of soil moisture. However, because mowing reduced soil vegetation coverage and litter accumulation, soil moisture and nutrient transported to the soil decreased. This would influence root growth, root respiration, soil microbial activity, and community composition. Secondly, mowing increased soil temperature [nearly 3°C higher than that at M3 in winter (p<0.01)]. Simultaneously, the CH₄ uptake decreased dramatically since CH₄ oxidation was likely to be limited due to low microbial activity with reduced soil temperature ( Liu et al., 2007; Zhang et al., 2012). In our study, we further found significant positive correlation between soil temperature and CH₄ uptake (p<0.01), which was consistent with that reported by Groffman et al. ( 2006). The R² values of the fitted regression curve indicate that the variation in soil temperature may explain 32%–34% (p<0.01) of the variation in CH₄ uptake (Figs. 8(a) and 8(b)). Other previous studies have also reported that soil temperature in the topmost layer might be the primary driving factor for CH₄ uptake ( Wang et al., 2005). Thirdly, we found soil \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N content at UM (means of 5.82 and 11.09 mg·kg^-1, respectively) were greater than those at M3 (means of 4.66 and 10.07 mg·kg^-1, respectively) (Table 3). The increase in CH₄ consumption may be attributable to the decrease in soil mineral N contents, since both \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N ( Steinkamp et al., 2001) and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N ( Kähkönen et al., 2002) could inhibit CH₄ oxidation. Whiting and Chanton (1993) also reported that reduction in inorganic N by mowing resulted in an increase of CH₄ oxidation. However, no significant difference (p>0.05) was found with respect to the soil \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N content because of restoration in three years, suggesting soil inorganic nitrogen content was not the leading factor in terms of the differences in CH₄ flux. There are some other factors that may affect the CH₄ uptake, such as the variation in root/shoot ratios ( Nitschke et al., 2010) and root exudation, favoring the microbial activity ( Lipson and Schmidt, 2004). The oxidation of atmospheric CH₄ by soils is mainly controlled by the activity of methanotrophic bacteria. Singh and Tate ( 2007) observed atmospheric CH₄ consumption in soil cores in New Zealand, for which type II methanotrophs may be chiefly responsible. We presumed that mowing increased CH₄ uptake perhaps due to changes in methanotrophic community structure and activity.

In this study, there was no correlation found between daily N₂O emission fluxes and simultaneously measured soil temperature and moisture in the UM steppe over the whole year (p>0.05), but significant correlation was found between daily N₂O emission flux and soil temperature in the M3 steppe (p<0.05). The R² values of the fitted regression curve indicate that the variation in soil temperature may explain 19%–20% (p<0.05) of the variation in N₂O emission (Figs. 8(c) and 8(d)). The annual range of temperature was wide (from -13°C to 26.5°C) in the mowing region, while the optimum temperature ranges for nitrification and denitrification were 15°C–35°C and 5°C–75°C, respectively ( Liu et al., 2010). Therefore, soil microbial activity was limited by temperature. We can thus conclude that soil moisture was not the key factor, but that soil temperature was the dominant factor accounting for the difference in N₂O emission between the UM (6.3 μg N·m^-2·h^-1) and M3 sites (4.3 μg N·m^-2·h^-1). Secondly, soil \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N (mean of 5.82 versus 4.66 mg·kg^-1) and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N (11.09 versus 10.07 mg·kg^-1) concentrations were higher in the UM than M3 fields. This may be explained by a faster relative mineral-N turnover in the mown fields. Li ( 1999) revealed that N mineralization by roots was accelerated as a result of increased \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N concentrations in the soil, especially during the maximum growth stage of grassland vegetation. In addition, differences in population density of individual plant communities was also found (Leymus chinense and Pulsatilla turczaninovii as the dominant species in M3 and UM, respectively) (data not shown), and so the change in availability of inorganic N also reflects a shift in plant community structure to some extent. Thirdly, N₂O is predominantly produced by microbial nitrification, which is affected by alterations of the soil environment in most semi-arid grasslands ( Verchot et al., 2002; Xu et al., 2008). Newton et al. ( 2014) reported that a change of the dominant species would affect the composition of soil microorganisms. We cannot exclude the effects of differences in SOM quality or competition processes between plants and microorganisms for N on nutrient availability between both sites, because such differences have been shown to also feed back on N₂O emissions ( Groffman et al. 2006), which is inversely related to soil inorganic nitrogen contents ( \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N) and methane uptake. The alteration of soil environment would affect the process of N₂O production and the related microbial functional groups. Therefore, further investigations are necessary.

Generally, the practice of mowing may have promoted CH₄ uptake and inhibited N₂O emissions through three pathways. First, mowing could reduce soil moisture content and increase soil temperature. Secondly, it could reduce soil inorganic N ( \mathrm{N}{\mathrm{H}}_{\mathrm{4}}^{+} -N and \mathrm{N}{\mathrm{O}}_{\mathrm{3}}^{-} -N) and SOM. And finally, a shift in plant species richness and soil microbial community composition could be affected by mowing.

Comparison with another grassland management practice

In the semiarid grasslands of Hulunber, Inner Mongolia, grazing is another important management practice. Up to now, published results about the effect of grazing on N₂O production are rather controversial. Some studies have found grazing could suppress N₂O production ( Wang et al., 2005; Xu et al., 2008; Wolf et al., 2010). Liu et al. (2010) presented results of 2 years (from January 2005 to December 2006) of measurements of N₂O fluxes from the native and grazed Leymus chinensis (LC) steppes in Inner Mongolia, and found the annual net emission of N₂O was 0.30 and 0.26 kg N₂O-N ha^-1 from the native and grazed LC steppes, respectively. Others, meanwhile, have reported that grazing could enhance N₂O production ( Frank et al., 2000; Saggar et al., 2004). Previous studies have reported that grazing tends to reduce CH₄ uptake in some grassland ecosystems. Liu et al. ( 2007) conducted CH₄ flux measurements during the periods of June to September 2004, May to September 2005, and March to June 2006, and concluded that the annual rate of CH₄ uptake was 3.58 and 1.91 kg C·ha^-1 at ungrazed and winter-grazed sites, respectively. Chen et al. ( 2010) noted that the annual CH₄ uptake was 3.7 and 2.1 kg C·ha^-1 during the year-round CH₄ flux measurements at an ungrazed site and winter-grazed site from June 2007 to October 2008. In contrast to grazing, mowing also affected the fluxes of CH₄ and N₂O in the meadow-steppe. Moreover, it had a larger potential to increase the CH₄ uptake and reduce N₂O emissions. As shown in Fig. 6 and Table 5, the CH₄ uptake was significantly higher at M3 (70.5 μg C·m^-2·h^-1) than at UM (40.3 μg C·m^-2·h^-1) (p<0.01); whereas, the N₂O emission was lower than that at UM (4.3 versus 6.3 μg N·m^-2·h^-1) (p<0.05) during the observation periods.

The area of natural grassland in China is 3.9 × 10⁸ ha. If all natural grassland behaved in the same way as the mown site in this study, the CO₂ equivalent of the annual net emission of CH₄ and N₂O from all natural grassland in China would be -40.3 × 10² Tg C·yr^-1 (Table 5). An estimation by neglecting the effect of mowing yields a figure of approximately 140.6 × 10² Tg C·yr^-1. If we have overestimated, as described above, at global and regional/national scales, then the sink and source are obviously reversed. Therefore, this is an important sink that cannot be neglected.

Conclusions

Our study has demonstrated that mowing could enhance CH₄ uptake and inhibit N₂O emissions in the semi-arid steppe ecosystem, mainly due to its effect on vegetation types and soil factors, especially on soil temperature. Soil moisture and inorganic nitrogen were not found to be the primary and key dominant factors. Therefore, further investigation is needed. Laboratory experiment results revealed that mowing changed the soil and environmental conditions that determine emissions of N₂O during thawing periods. Our results suggest that the practice of mowing could mitigate potential global warming in terms of CH₄ uptake and N₂O emissions, and mown steppes may be a neglected sink for GHGs.

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