Grazing offsets the stimulating effects of nitrogen addition on soil CH4 emissions in a meadow steppe in Northeast China
Authors:
Rongrong Ren aff001; Wanling Xu aff001; Mingming Zhao aff001; Wei Sun aff001
Authors place of work:
Key Laboratory for Vegetation Ecology, Ministry of Education Institute of Grassland Science, Northeast Normal University, Changchun, Jilin Province, China
aff001
Published in the journal:
PLoS ONE 14(12)
Category:
Research Article
doi:
https://doi.org/10.1371/journal.pone.0225862
Summary
Grazing is the most common land use type for grasslands, and grazing may alter the impacts of the predicted enhancement of nitrogen deposition on soil CH4 flux. To understand the effects of nitrogen addition, grazing, and their interactions on soil CH4 flux, we conducted a field study on CH4 flux in a meadow steppe in Northeast China from 2017 to 2018. We measured the soil CH4 flux and soil physiochemical and vegetation parameters. The studied meadow steppe soil acted as a CH4 source due to the legacy effects of an extreme rainfall event. During the experimental period, the average CH4 fluxes were 7.8 ± 1.0, 5.8 ± 0.5, 9.3 ± 0.9 and 7.6 ± 0.6 μg m-2 h-1 for the CK (control), G (grazing), N (nitrogen addition) and NG (grazing and nitrogen addition) treatments, respectively. The cumulative CH4 fluxes were 24.9 ± 2.6, 11.5 ± 4.9, 28.8 ± 4.2 and 17.8 ± 3.5 μg m-2 yr-1 for the CK, G, N and NG treatments, respectively. The N addition increased the average CH4 flux by 19%, and the grazing treatment reduced it by 25%. The soil CH4 flux was positively correlated with the 0–10 cm soil water filled pore space (P < 0.01), soil NH4+-N (P < 0.01) and soil NO3--N (P < 0.01), but negatively correlated with the 0–10 cm soil temperature (P < 0.01), except for the sampling dates that were strongly influenced by the extreme rainfall event. The average CH4 flux was significantly (P < 0.05) affected by the grazing and N addition treatments with the N addition treatment significantly (P < 0.05) increased the CH4 flux, whereas grazing significantly (P < 0.05) decreased the CH4 flux. Grazing offset the stimulating effects of N addition on CH4 flux, and there was no difference (P = 0.79) in the CH4 flux between the CK and NG plots. In summary, moderate grazing has the potential to reduce the negative impacts of N addition on CH4 flux and can increase the capacity of the soil CH4 sink in the studied meadow steppe.
Keywords:
Cattle – Edaphology – Biomass – Rain – Grasslands – Grazing – Oxidation – Soil pH
Introduction
As the second most important greenhouse gas, methane (CH4) has a global warming potential 34 times that of carbon dioxide (CO2) for the time horizon of 100 years and contributes approximately 25% to global warming [1]. The atmospheric CH4 concentration has increased from 722 ppb to 1803 ppb since the industrial revolution [1]. Soil CH4 flux is determined by the balance between methanogenic bacteria-associated CH4 production and methanotrophic bacteria-related CH4 consumption [2]. Other than soil microbial activity, CH4 flux is also influenced by soil physiochemical conditions, such as moisture, temperature, pH, organic C content and inorganic N content [3,4]. In an African tropical montane region, the combined contribution of soil water content and bulk density explained over 49% of the total variation in soil CH4 fluxes [5]. These abiotic and biotic factors are likely to be affected by global change factors (e.g., nitrogen addition, altered precipitation regime) as well as land use patterns and extent (e.g., grazing), and subsequently modify soil CH4 flux [6,7]. Indeed, there is evidence that CH4 uptake differs significantly among land-cover types, suggesting that CH4 uptake may respond differently in different land-covers or land-use change scenarios [8].
Nitrogen (N) deposition has continued to increase since the last century, mainly due to fossil fuel combustion [9]. The impacts of enhanced N input on CH4 flux have attracted much attention, but the results are not consistent [10–12]. The response of soil CH4 flux to N addition is still uncertain and is mainly determined by the forms and rates of N addition and soil properties [13]. The results of a meta-analysis showed that N addition reduced CH4 uptake by 38% [13]. It has often been reported that CH4 uptake is significantly reduced by long-term N addition, which can change soil pH and nitrate N content [14]. There is evidence that long-term N addition may increase the availability of soil NH4+ and soil NO3-, thus changing the activity of methanogenic bacteria and methanotrophic bacteria, and consequently affecting soil CH4 flux [10]. However, some studies have suggested that the content of soil NH4+, but not NO3-, is the dominant factor controlling CH4 oxidation rates [15]. High soil NH4+ concentrations may reduce the oxidation capacity of soils for atmospheric CH4, thus decreasing the net flux of CH4 from the atmosphere to the soil [15]. N addition can also promote plant growth and trigger changes in soil properties, such as soil moisture and soil temperature, which may also affect CH4 flux. In addition to the uncertainty concerning the effects of N addition on CH4 flux, land use patterns, such as grazing, may also interact with N addition and alter CH4 flux.
Grazing is the major land use type for grasslands, and its impact on the production and consumption of CH4 flux has been reported widely [16–18]. Grazing alters CH4 flux by affecting plant productivity, soil properties and soil microbial activity [19–21]. First, livestock trampling changes the soil bulk density and weakens the diffusivity of CH4 and O2 across the soil profile [22]. Second, grazing can reduce the aboveground biomass [23] and litter biomass, which can affect soil temperature and soil moisture. Third, grazing may alter the composition and activity of methanogenic bacteria and methanotrophic bacteria [24]. The impacts of grazing on CH4 highly depend on grazing intensity, grazing time and grazing site [18]. Grazing may offset N deposition-associated negative effects on CH4 flux by changing soil N availability and soil physical characteristics. However, little information is available regarding the interactive effects of N addition and grazing on soil CH4 flux.
Simulated atmospheric N deposition and/or grazing experiments have been conducted in many natural ecosystems. However, the impact of N deposition and grazing on the production or consumption of CH4 are not well understood. To understand the effects of N addition, grazing, and their interactions on grassland CH4 flux, we conducted a field study on CH4 flux in a meadow steppe in Northeast China. The main aims of this study were (1) to investigate the seasonal dynamics of CH4 flux and unravel the underlying mechanisms in the meadow steppe; (2) to examine the effects of N addition, grazing and their interactions on soil CH4 flux. We hypothesized that (1) the soil CH4 flux in the studied meadow steppe would display strong seasonal variation and may be highly influenced by the WFPS (water filled pore space), topsoil temperature and inorganic N content; (2) the N addition treatment would increase the CH4 flux, whereas grazing would decrease the CH4 flux and the grazing treatment would offsets the stimulating effects of N addition on the CH4 flux.
Materials and methods
Ethics statement
No specific permissions were required to conduct research at the field site, because the Songnen Grassland Ecological Research Station is a department of the Northeast Normal University. No specific permissions were required for the study either, as it was conducted in accordance with the guidelines set by the Northeast Normal University. No specific permissions were required for the locations or the activities. No location was privately owned or protected in any way, and the field studies did not involve endangered or protected species. The cattle used on the experiment were rented from a ranch. During the experimental grazing period, the cattle had access to food and water. After the experimental grazing, the cattle grazed freely in the surrounding grasslands.
Site description
This experiment was conducted in the Songnen meadow steppe, which is located in western Jilin Province, Northeast China (44°40′-44°44′ N, 123°44′-123°47′ E). The study area is influenced by a temperate semiarid monsoon climate. The annual average temperature is 6.4°C (1950–2004), and the frost-free period is 150 days. The average annual precipitation is 471 mm (1950–2004), with over 70% occurring from June to August [25]. The vegetation is dominated by Leymus chinensis, a C3 perennial rhizomatous grass. Phragmites australis (a perennial C3 plant), Chloris virgata (an annual C4 plant) and Kalimeris integrifolia (a perennial C3 plant) are also abundant [26]. The vegetation coverage ranged from 50% to 90%, with 100–360 g m-2 of aboveground biomass in the peak biomass season [27]. The soil in the study area is saline-alkaline, which is equivalent to an Aqui-Alkalic Halosol based on the Chinese soil classification or a Salic Solonetz in the World Reference Base for Soil Resources (WRB) [28]. The studied soil had an organic carbon content of 2.0% and a total nitrogen content of 0.15%, and the soil pH ranged from 8.0 to 9.0 [29]. The available soil phosphorus content was 2.5 mg kg- 1. The soils contained high contents of free sodium bicarbonate (NaHCO3) and sodium carbonate (Na2CO3) [27].
Experimental design
In 2010, we fenced a grassland with an area of 400 m × 100 m in the experimental site. Grazing by large herbivores and mowing were excluded from the fenced area. In August 2015, 4 blocks (100 m × 100 m each) with similar vegetation compositions were established within the fenced grassland. We laid out 4 plots (30 m × 30 m each) in each block. There were no significant differences in the vegetation characteristics among the blocks and plots. Within each block, we randomly assigned one of four treatments (control, CK; grazing, G; nitrogen addition, N; and grazing and nitrogen addition, NG) to each plot. The grazing and nitrogen addition treatments were initiated in May 2016. For each month of the growing season (May to September), each of the G and NG plots received one-day (06:00 AM to 10:00 AM) moderate grazing (approximately 50% of the aboveground biomass was consumed by herbivores) by adult Simmental cattle [30]. To minimize stocking rate differences between the G and NG treatments, 14 cattle and 18 cattle were allowed to graze in the G and NG plots, respectively. This assignment greatly reduced stocking rate differences between the G and NG treatments. One week after the grazing treatment, urea (2 g m-2) was manually spread on the N and NG plots. For each year of the experiment, urea was applied five times (once per month from May to September), which resulted in an N addition rate of 10 g m-2 yr-1.
Sampling and measurement of CH4 flux
Soil CH4 fluxes were measured from the beginning of April 2017 through the end of September 2018. Gas sampling was conducted once per week during the growing season (May-September), twice a month during the non-growing season (April, October, and November), once a month during the winter (December, January, and February), and every three days during the freezing and thawing period (March).
The static opaque chamber method was used for the measurements of soil CH4 flux [24]. The static chamber consisted of two parts: a stainless steel (length × width × height = 30 cm × 30 cm × 15 cm) base and a box (length × width × height = 30 cm × 30 cm × 60 cm) made of polypropylene. To prevent the influence of direct radiative heating during the sampling period, the outside of the chamber was covered by reflective aluminium foil. A fan was installed inside the chamber to mix the air. The stainless steel base had a groove to connect with the chamber, and during the gas sampling period, the chamber was sealed by the addition of water to the groove. The steel base was inserted into the soil surface (10 cm) before the gas sampling.
During each gas sampling campaign, gas samples were collected between 09:00 AM and 11:00 AM (China Standard Time, CST) [31]. The chambers were closed for an hour, and gas samples (200 ml) were collected every 20 min using plastic syringes. The temperature inside the chamber was recorded at the same time using an electronic thermometer (DT-1, Jingchuang, China). The collected gas samples were stored in previously evacuated gas sampling bags (200 ml) before the laboratory CH4 concentration measurements. In each plot, we randomly selected six points for gas sampling and sampled these points during each of the subsequent gas sampling campaigns. The CH4 concentration was determined using a CH4/N2O gas analyser (Model 913–1054, Los Gatos Research, USA) within one week of field sampling.
CH4 flux was calculated as follows:
Soil temperature and soil water filled pore space
Soil temperature and soil water-filled pore space (WFPS) at a depth of 0–10 cm were monitored during each gas sampling event near the gas sampling points. Soil temperature was measured by a soil temperature probe (TPG-21, Tuopu, China) with six replicates in each plot.
For the measurement of WFPS, we used the ring knife method to collect the soil samples (50 cm-3). The collected soil samples were placed in aluminium boxes, and the fresh weight was measured. The dry weight was measured after oven-drying at 105°C to a constant weight [34]. The volumetric soil moisture was calculated as the water loss divided by the soil core volume (50 cm-3). The soil bulk density was calculated by the ratio between the dry weight and the soil core volume. Based on the volumetric soil moisture, soil bulk density, and a particle size density of 2.65 g cm-3 [35], the WFPS was calculated as follows:
Precipitation
Precipitation was measured using an RG2-M sensor (Onset Computer Corporation, Bourne, MA, USA) 0.2 km distant from the experimental site.
Soil sampling
Using a soil auger (2.5 cm in diameter), soil samples (0–10 cm) were collected near the gas sampling point once per month from April 2017 to September 2018. For each gas sampling point, we sampled soils from 6 locations and thoroughly mixed them together into a combined sample during each soil sampling campaign. The mixed soils were then sieved (2 mm) to remove roots and stones in the field and kept in a refrigerator at -20°C before analysis.
Soil NH4+-N, soil NO3--N and soil pH
Ten grams of moist soil were placed in an Erlenmeyer flask and extracted with 50 ml KCl (2 M) and shaken for 1 h. After shaking, the samples were allowed to settle and were filtered through filter paper; then, the samples were stored at -20°C before analysis. The soil NH4+-N and soil NO3—N contents were measured by a continuous flow autoanalyser (FUTURA, AMS Alliance, Italia). Soil pH was measured by mixing 10 g soil with water in a 1:5 ratio. The resulting slurry was stirred for 1 h and then allowed to sit for another hour before pH was measured using a pH probe (Phs-3c, Shanghai, China) [36].
Vegetation survey and biomass measurement
A vegetation survey and biomass measurements were conducted once per year in 2017 and 2018. In each plot, the number of plant species and the number of individuals of each species were counted in six randomly placed quadrats (50 cm × 50 cm) at the end of July (approximately two weeks after the grazing treatment in July). The aboveground biomass (AGB) was harvested after the vegetation survey. Litter was also collected. The belowground biomass (BGB) was determined by washing roots out of a soil core with a diameter of 7.0 cm collected at a depth of 0–10 cm [37]. One soil core randomly collected in each quadrat was used for the vegetation survey. The AGB, BGB and litter samples were oven-dried at 65°C to a constant weight (approximately 48 h).
Statistical analysis
The effects of G, N and their interactions on soil temperature, soil WFPS, soil inorganic N (NH4+-N and NO3--N) content, average CH4 flux and cumulative CH4 flux were assessed with two-way ANOVAs. The influences of G, N, year and their interactions on the average CH4 flux, soil temperature, soil WFPS and biomass (AGB, BGB, and litter) were examined using three-way ANOVAs. Statistical significance was set at P < 0.05. Linear regression analyses were used to examine the dependence of CH4 flux on soil temperature, soil WFPS, soil NH4+-N, soil NO3--N, soil pH and plant biomass. All statistical analyses were conducted using SPSS 22.0 software (SPSS, Inc., Chicago, USA), and the graphs were generated using SigmaPlot 12.5 software (Systat Software, Inc., Chicago, USA). The results were expressed as the mean value ± 1 standard error (n = 4).
Results
Environmental variables
Over the experimental period, precipitation exhibited strong variation in timing and event size (Fig 1A). For example, there was an extreme rainfall event (203 mm) in August 2017 that accounted for 49% of the growing season precipitation. The amount of growing season precipitation (April—October) in 2018 (308 mm) was lower than that in 2017 (412 mm) (Fig 1A). The air temperature had clear seasonal patterns, being highest (31°C) in July and August and lowest (-29°C) in January (Fig 1A). The seasonality of the 0–10 cm soil temperature coincided with the seasonal patterns of air temperature. For both experimental years, the 0–10 cm soil temperature peaked (28°C) at the end of July (Fig 1B). During the growing season, the 0–10 cm soil WFPS varied substantially and was closely related to the precipitation events (Fig 1C).
CH4 flux
The CH4 flux varied substantially during the experimental period, and there were no apparent differences in the variation patterns among the treatments (Fig 2A). From April to July 2017, CH4 fluxes were negative for most of the sampling dates, which indicated the studied grassland was a net CH4 sink during that time period. Following the extreme precipitation event (August 2017, 203 mm), there was a strong increase in CH4 emissions for all treatments, with the maximum CH4 emission rates (241.2 ± 8.1, 245.6 ± 4.3, 260.2 ± 24.6 and 253.6 ± 11.6 μg m-2 h-1 for the CK, G, N and NG treatments, respectively) were observed on 10 September. Then, the CH4 emission rate gradually decreased and tended to stabilize until March 2018. From April to September in 2018, the CH4 uptake rates in the studied grassland were low for most of the sampling dates; however, pulses of CH4 emissions were also observed occasionally (Fig 2A).
The monthly average CH4 fluxes were negative from April to July in 2017 and were positive from August 2017 to March 2018. The monthly average CH4 fluxes changed to negative after March 2018, and the studied grassland was a net CH4 sink from April to August 2018 (Fig 2B). During the experimental period, the average CH4 fluxes were 7.8 ± 1.0, 5.8 ± 0.5, 9.3 ± 0.9 and 7.6 ± 0.6 μg m-2 h-1 for the CK, G, N and NG treatments, respectively. The average CH4 flux was significantly (P < 0.05) affected by the grazing and N addition treatments. The nitrogen addition increased the average CH4 flux by 19%, and the grazing treatment reduced it by 25%. There were no significant interactive effects between the grazing and N addition treatments on average CH4 flux (P = 0.79) (Fig 2C).
The cumulative CH4 fluxes were 24.9 ± 2.6, 11.5 ± 4.9, 28.8 ± 4.2 and 17.8 ± 3.5 μg m-2 yr-1 for the CK, G, N and NG treatments, respectively. The cumulative CH4 flux was reduced significantly in the grazed plots (P < 0.01); however, it was not affected (P = 0.21) by the N addition. For all treatments, the cumulative CH4 fluxes (October 2017 to September 2018) were positive, indicating that CH4 emissions were higher than uptake (Fig 2D). The highest cumulative CH4 flux (October 2017 to September 2018) was observed in the N plots (28.8 ± 4.2 mg m-2 yr-1), whereas the lowest value was detected in the grazing plots (11.5 ± 4.9 mg yr-1).
Aboveground biomass, belowground biomass and litter biomass
Aboveground biomass (AGB) was stimulated by the N addition (Fig 3A; S1 Table); the AGB was enhanced by 41.9% and 50.3% in 2017 and 2018, respectively, relative to the control treatment. In contrast, grazing significantly (P < 0.01) reduced the AGB. The two studied growing seasons differed significantly (P < 0.01) in AGB, with it was higher in 2018 than in 2017. Only the N addition had a significant impact on belowground biomass (BGB) (Fig 3B; S1 Table). Grazing significantly (P < 0.01) reduced litter mass, while the N addition significantly (P < 0.01) increased litter mass (Fig 3C; S1 Table). The litter mass was higher in 2018 than in 2017.
Effects of grazing and nitrogen addition on the soil physiochemical parameters
Compared to the CK plots, the grazing treatment significantly (P < 0.01) increased the soil temperature by 1.5°C, whereas the N addition treatment significantly (P < 0.01) reduced the soil temperature by 1.25°C (Fig 4A). In contrast, the grazing treatment significantly (P < 0.01) reduced the soil WFPS by 10.6%, and N addition significantly (P < 0.01) increased it by 9.5% (Fig 4B) relative to the CK treatment. There were no significant interactive effects between grazing and N addition on the soil temperature and soil WFPS (Fig 4A & 4B).
Compared with the CK treatment, N addition significantly (P < 0.01) enhanced soil NH4+-N and soil NO3--N contents, whereas the grazing treatment had no significant effects on soil NH4+-N and NO3--N contents. There were significant (P < 0.05) interactive effects between grazing and N addition on soil NO3--N content, but not on soil NH4+-N content (Fig 4C & 4D).
Dependence of CH4 flux on plant biomass and soil parameters
Average CH4 fluxes were positively correlated with aboveground biomass (R2 = 0.34, P = 0.02) (Fig 5A). However, the relationships between average CH4 flux and belowground biomass (R2 = 0.02, P = 0.64) and litter mass (R2 = 0.12, P = 0.19) in 2018 (Fig 5B & 5C) were not statistically significant. The soil CH4 fluxes were positively correlated with the soil WFPS (R2 = 0.71–0.88, P < 0.01) at 0–10 cm depth and negatively correlated with the soil temperature (R2 = 0.34–0.52, P < 0.01) at 0–10 cm depth from 2017 to 2018; the exceptions were the sampling dates that were strongly influenced by the extreme precipitation event (Fig 6A & 6B). CH4 fluxes were positively correlated with soil NH4+-N (R2 = 0.26, P < 0.01) and soil NO3—N (R2 = 0.16, P < 0.01) (Fig 6C & 6D). There was no significant (P = 0.10) relationship between soil pH at 0–10 cm depth and average CH4 flux (Fig 6E).
Discussion
Soil CH4 fluxes
Although CH4 uptake (negative flux) was occasionally found in all treatments, the average CH4 flux and cumulative CH4 flux were positive, indicating that our study site acted as a net source for atmospheric CH4 during the experimental period (Fig 2A–2D). This result contrasts with the results of most studies in grasslands in China [38]. The average CH4 fluxes in the four treatments ranged from 5.8 μg m-2 h-1 to 9.3 μg m-2 h-1 (Fig 2C), which were significantly higher than fluxes reported in typical grasslands in Inner Mongolia [39] and alpine steppe grasslands on the Tibetan Plateau [40]. This inconsistency may be attributed to the occurrence of an extreme precipitation event that occurred in August 2017. The extreme precipitation event strongly altered the soil WFPS and soil temperature, which promoted methanogenic activity and shifted the ecosystem from a CH4 sink to a source. In a previous study, Petrakis et al. (2017) reported that extreme inundation changed an ecosystem from a small CH4 sink to a small source of CH4 to the atmosphere [41], which highlights the importance of soil water content in controlling the balance of soil CH4 production and consumption [42]. Our observations showed that the potential of meadow steppe soils to act as CH4 sinks may decrease after extreme precipitation events.
Effects of nitrogen addition on CH4 flux
Our results showed that N addition significantly increased soil CH4 emissions (Fig 2B & 2C), which is consistent with the results of most previous studies [11,43]. However, no effects of N addition on CH4 exchange were reported in subtropical plantation soils [44] or in degraded steppe soils [45]. N addition may influence CH4 flux by decreasing the uptake of CH4 or by increasing CH4 production. N addition enhanced the contents of soil NH4+-N and NO3—N, which are strong inhibitors of CH4 uptake [44]. First, NH4+ is a competitive inhibitor of CH4 oxidation due to the lack of specificity of methane monooxygenase (MMO) in methanotrophs [46]. Second, NH4+ oxidizes to the intermediate hydroxylamine (NH2OH) by CH4 monooxygenase or is further oxidized by other enzyme systems of methanotrophs to the end product of nitrite (NO2-). NO2- is also produced via NO3- reduction in anaerobic microsites. Hydroxylamine (NH2OH) and nitrite (NO2-) are toxic to methanotrophic bacteria [47]. Third, osmotic stress caused by the added nitrogen salt can suppress the activity of methanotrophs [48]. This inference was supported by the positive effects of N addition on the soil NH4+-N and NO3—N contents (Fig 4C & 4D) and the positive dependence of CH4 flux on the soil NH4+-N and NO3-N contents (Fig 6C & 6D). A previously published study suggested that CH4 uptake rate was sensitive to changes in pH and soil NO3—N, rather than NH4+-N [14]. However, we only detected a marginally significant relationship between soil pH and average CH4 flux (Fig 6E).
In addition to the effects on CH4 oxidation, N addition promoted biomass production and litter input (Fig 3C), which subsequently alleviated microbial C limitation [49]. As a result, the activities of methanogenic archaea were enhanced, and more CH4 was produced [49]. Moreover, we found average CH4 fluxes were positively correlated with aboveground biomass (Fig 5A). This likely occurred because increased aboveground biomass (i.e. greater canopy cover) reduces soil temperature via increased soil shading [50], which further decreases soil evaporation and increases soil moisture [51]. Grazing reduced the aboveground biomass while nitrogen addition increased aboveground biomass, so grazing increased soil temperature while nitrogen addition reduced soil temperature. This inference was supported by the positive dependence of soil CH4 flux on soil WFPS, and the negative correlation between soil CH4 flux and soil temperature.
Effects of grazing on CH4 fluxes
Grazing is the most important human practice in grasslands and plays an important role in regulating the emission and uptake of greenhouse gases [17,18]. Grazing influences CH4 flux by altering aboveground and belowground productivity, litter input, microbial composition and soil physiochemical properties (soil temperature, soil moisture, soil nutrient content, etc.) but grazing effects depend highly on grazing intensity and duration. A previous study reported that grazing experiments with durations longer than 5 years had a significant effect on soil CH4 uptake, while experiments with durations less than 5 years had no effect on soil CH4 uptake [18]. In our study, the G and NG plots received moderate grazing by cattle in the growing season from 2016 to 2018, which was less than 5 years. However, grazing significantly decreased the average CH4 fluxes in our study (Fig 2B & 2C), which suggests that the CH4 fluxes in the studied meadow steppe are very sensitive to grazing disturbance.
Grazing is associated with shifts in plant biomass and soil structure and changes in soil temperature and soil moisture, all of which likely affect soil CH4 flux. First, grazing can alter soil temperature by increasing the radiant energy that reaches it, leading to higher soil temperature. Grazing reduced the surface vegetation coverage and litter biomass (Fig 3A & 3C) and enhanced soil temperature (Fig 4A), which likely had a greater impact on CH4 oxidation than on methanogen activities [32]. This inference was supported by the negative correlation between soil CH4 flux and soil temperature (Fig 6B). Methanotrophs are highly temperature sensitive [52], such that the rate at which they consume CH4 increases (i.e., a more negative flux rate) with increasing temperature [53]. Second, grazing decreased soil moisture (Fig 4B) by enhancing the soil temperature (Fig 4A) and evaporation [23]. Soil moisture influences CH4 flux by affecting microbial activities and influencing CH4 diffusion [40]. Grazing-induced reductions in soil moisture favour the diffusion of CH4 to methanotrophs in the subsurface soil, therefore increasing CH4 consumption. In line with previous work [54], we detected a strong positive dependence of soil CH4 fluxes on the soil WFPS at 0–10 cm depth (Fig 6A). Third, animal trampling disturbs the topsoil and decreases the diffusion of CH4 and oxygen from the atmosphere into the soil profile [55] or from the soil into the atmosphere, which can directly affect the soil CH4 flux. In the present study, the grazing treatment was carried out for only 3 years; therefore, trampling was unlikely to have significant impacts on soil compaction. Finally, decreases in aboveground biomass and litter mass production by grazing and the corresponding decline in C that is available as a substrate in the soil often result in lower soil CH4 emissions [56], which can lead to a decline in CH4 flux.
Interactive effects of nitrogen addition and grazing on CH4 fluxes
Although N addition significantly increased the CH4 flux and grazing significantly decreased the CH4 flux, there were no significant interactive effects between N addition and grazing on CH4 flux during the experimental period (S1 Table). In general, N addition and grazing had opposite effects on the soil environmental conditions and vegetation parameters (Fig 4A & 4B). For example, soil temperature decreased in the N addition plots, whereas soil temperature increased in the grazing plots. Moreover, N addition increased AGB and litter, but grazing had the opposite effect. Therefore, when N addition and grazing were combined, no significant differences were detected between the NG and CK treatments for the aforementioned soil and vegetation parameters or for CH4 flux. Our results suggest that grazing management has the potential to offset the stimulating effects of N addition on CH4 emissions, which highlights the importance of land management for the estimation of global change factor effects on CH4 flux.
Conclusions
Soil CH4 flux in the studied meadow steppe displayed strong seasonal variation and was highly influenced by the WFPS, topsoil temperature and inorganic N content. Due to the occurrence of an extreme rainfall event, the studied ecosystem was a net CH4 source during the experimental period. The N addition treatment significantly increased CH4 flux, whereas grazing significantly decreased CH4 flux. Grazing offset the stimulating effects of N addition on CH4 flux, and there was no difference in CH4 flux between the CK and NG treatments. Our results suggest that moderate grazing has the potential to reduce the negative impacts of N addition on CH4 flux and can increase the capacity of the soil CH4 sink in the studied meadow steppe. The present study highlights the importance of grassland management on the regulation of the response of ecosystem processes to global change stresses.
Supporting information
Zdroje
1. Stocker TF, Qin D, Plattner GK, Tignor MMB, Allen SK, Boschung J, et al. (2013) Climate change 2013 the physical science basis: Working Group I contribution to the fifth assessment report of the intergovernmental panel on climate change. 1–1535 p.
2. Tate KR (2015) Soil methane oxidation and land-use change—from process to mitigation. Soil Biology and Biochemistry 80: 260–272.
3. Lilly A, Ball BC, McTaggart IP, Degroote J (2009) Spatial modelling of nitrous oxide emissions at the national scale using soil, climate and land use information. Global Change Biology 15: 2321–2332.
4. Smith KA, Ball T, Conen F, Dobbie KE, Massheder J, Rey A (2003) Exchange of greenhouse gases between soil and atmosphere: Interactions of soil physical factors and biological processes. European Journal of Soil Science 54: 779–791.
5. Wanyama I, Pelster DE, Butterbach-Bahl K, Verchot LV, Martius C, Rufino MC. (2019) Soil carbon dioxide and methane fluxes from forests and other land use types in an African tropical montane region. Biogeochemistry 143: 171–190.
6. Guetlein A, Gerschlauer F, Kikoti I, Kiese R (2018) Impacts of climate and land use on N2O and CH4 fluxes from tropical ecosystems in the Mt. Kilimanjaro region, Tanzania. Global Change Biology 24: 1239–1255. doi: 10.1111/gcb.13944 29044840
7. Zhou M, Zhu B, Brueggemann N, Bergmann J, Wang Y, Butterbach-Bahl K (2014) N2O and CH4 emissions, and NO3- leaching on a crop-yield basis from a subtropical rain-fed wheat-maize rotation in response to different types of nitrogen fertilizer. Ecosystems 17: 286–301.
8. Li J, Nie M, Pendall E (2019) An incubation study of temperature sensitivity of greenhouse gas fluxes in three land-cover types near Sydney, Australia. Science of The Total Environment 688: 324–332. doi: 10.1016/j.scitotenv.2019.06.206 31233914
9. Liu L, Zhang X, Lu X (2016) The composition, seasonal variation, and potential sources of the atmospheric wet sulfur (S) and nitrogen (N) deposition in the southwest of China. Environmental Science and Pollution Research 23: 6363–6375. doi: 10.1007/s11356-015-5844-1 26620861
10. Jang I, Lee S, Zoh KD, Kang H (2011) Methane concentrations and methanotrophic community structure influence the response of soil methane oxidation to nitrogen content in a temperate forest. Soil Biology and Biochemistry 43: 620–627.
11. Mosier AR, Parton WJ, Phongpan S (1998) Long-term large N and immediate small N addition effects on trace gas fluxes in the colorado shortgrass steppe. Biology and Fertility of Soils 28: 44–50.
12. Zhuang Q, Chen M, Xu K, Tang J, Saikawa E, Lu Y, et al. (2013) Response of global soil consumption of atmospheric methane to changes in atmospheric climate and nitrogen deposition. Global Biogeochemical Cycles 27: 650–663.
13. Liu L, Greaver TL (2009) A review of nitrogen enrichment effects on three biogenic GHGs: the CO2 sink may be largely offset by stimulated N2O and CH4 emission. Ecology Letters 12: 1103–1117. doi: 10.1111/j.1461-0248.2009.01351.x 19694782
14. Chen S, Hao T, Goulding K, Misselbrook T, Liu X (2019) Impact of 13-years of nitrogen addition on nitrous oxide and methane fluxes and ecosystem respiration in a temperate grassland. Environmental Pollution 252: 675–681. doi: 10.1016/j.envpol.2019.03.069 31185356
15. Kim YS, Imori M, Watanabe M, Hatano R, Yi MJ, Koike T (2012) Simulated nitrogen inputs influence methane and nitrous oxide fluxes from a young larch plantation in northern Japan. Atmospheric Environment 46: 36–44.
16. Schönbach P, Wolf B, Dickhöfer U, Wiesmeier M, Chen W, Wan H, et al. (2012) Grazing effects on the greenhouse gas balance of a temperate steppe ecosystem. Nutrient Cycling in Agroecosystems 93: 357–371.
17. Skiba U, Jones SK, Drewer J, Helfter C, Anderson M, Dinsmore K, et al. (2013) Comparison of soil greenhouse gas fluxes from extensive and intensive grazing in a temperate maritime climate. Biogeosciences 10: 1231–1241.
18. Tang S, Zhang Y, Zhai X, Wilkes A, Wang C, Wang K. (2018) Effect of grazing on methane uptake from Eurasian steppe of China. BMC Ecology 18 (1): 11. doi: 10.1186/s12898-018-0168-x 29558936
19. Cui X, Wang Y, Niu H, Wu J, Wang S, Schnug E, et al. (2005) Effect of long-term grazing on soil organic carbon content in semiarid steppes in Inner Mongolia. Ecological Research 20: 519–527.
20. Frank DA, Gehring CA, Machut L, Phillips M (2003) Soil community composition and the regulation of grazed temperate grassland. Oecologia 137: 603–609. doi: 10.1007/s00442-003-1385-2 14513350
21. Gao YZ, Giese M, Lin S, Sattelmacher B, Zhao Y, Brueck H (2008) Belowground net primary productivity and biomass allocation of a grassland in Inner Mongolia is affected by grazing intensity. Plant and Soil 307: 41–50.
22. Liu C, Holst J, Brüggemann N, Butterbach-Bahl K, Yao Z, Yue J, et al. (2007) Winter-grazing reduces methane uptake by soils of a typical semi-arid steppe in Inner Mongolia, China. Atmospheric Environment 41: 5948–5958.
23. Yates CJ, Norton DA, Hobbs RJ (2000) Grazing effects on plant cover, soil and microclimate in fragmented woodlands in south-western Australia: Implications for restoration. Austral Ecology 25: 36–47.
24. Zhou XQ, Wang YF, Huang XZ, Tian JQ, Hao YB (2008) Effect of grazing intensities on the activity and community structure of methane-oxidizing bacteria of grassland soil in Inner Mongolia. Nutrient Cycling in Agroecosystems 80: 145–152.
25. Cui H, Wang Y, Jiang Q, Chen S, Ma JY, Sun W. (2015) Carbon isotope composition of nighttime leaf-respired CO2 in the agricultural-pastoral zone of the songnen plain, northeast China. PLoS ONE 10 (9).
26. Wang Y, Jiang Q, Yang Z, Sun W, Wang D (2015) Effects of water and nitrogen addition on ecosystem carbon exchange in a meadow steppe. PLoS ONE 10 (5).
27. Yu P, Liu S, Xu Q, Fan G, Huang Y, Zhou DW (2019) Response of soil nutrients and stoichiometric ratios to short-term land use conversions in a salt-affected region, northeastern China. Ecological Engineering 129: 22–28.
28. Spaargaren O C., Deckers J (1998) The World Reference Base for Soil Resources. pp. 21–28.
29. Shi B, Wang Y, Meng B, Zhong S, Sun W (2018) Effects of nitrogen addition on the drought susceptibility of the Leymus chinensis meadow ecosystem vary with drought duration. Frontiers in Plant Science 9.
30. Shi B, Xu W, Zhu Y, Wang C, Loik ME, Sun W. (2019) Heterogeneity of grassland soil respiration: Antagonistic effects of grazing and nitrogen addition. Agricultural and Forest Meteorology 268: 215–223.
31. Xu M, Cheng S, Fang H, Yu G, Gao W, Wang YS, et al. (2014) Low-level nitrogen addition promotes net methane uptake in a boreal forest across the Great Xing'an Mountain Region, China. Forest Science 60: 973–981.
32. Wang X, Zhang Y, Huang D, Li Z, Zhang X (2015) Methane uptake and emissions in a typical steppe grazing system during the grazing season. Atmospheric Environment 105: 14–21.
33. Jiang C, Yu G, Fang H, Cao G, Li Y (2010) Short-term effect of increasing nitrogen deposition on CO2, CH4 and N2O fluxes in an alpine meadow on the Qinghai-Tibetan Plateau, China. Atmospheric Environment 44: 2920–2926.
34. Bai W, Fang Y, Zhou M, Xie T, Li L, Zhang WH (2015) Heavily intensified grazing reduces root production in an Inner Mongolia temperate steppe. Agriculture Ecosystems & Environment 200: 143–150.
35. Cardoso AS, Brito LF, Janusckiewicz ER, Morgado ES, Barbero RP, Koscheck JFW, et al. (2017) Impact of grazing intensity and seasons on greenhouse gas emissions in tropical grassland. Ecosystems 20: 845–859.
36. Courtois EA, Stahl C, Van den Berge J, Bréchet L, Van Langenhove L, Richter A, et al. (2018) Spatial variation of soil CO2, CH4 and N2O fluxes across topographical positions in tropical forests of the Guiana Shield. Ecosystems 21: 1445–1458.
37. Liu M, Liu G, Gong L, Wang D, Sun J (2014) Relationships of biomass with environmental factors in the grassland area of Hulunbuir, China. PLoS ONE 9.
38. Wang Y, Chen H, Zhu Q, Peng C, Wu N, Yang G, et al. (2014) Soil methane uptake by grasslands and forests in China. Soil Biology and Biochemistry 74: 70–81.
39. Wang Y, Xue M, Zheng X, Ji B, Du R, Wang YF (2005) Effects of environmental factors on N2O emission from and CH4 uptake by the typical grasslands in the Inner Mongolia. Chemosphere 58: 205–215. doi: 10.1016/j.chemosphere.2004.04.043 15571752
40. Li Y, Dong S, Liu S, Zhou H, Gao Q, Cao G, et al. (2015) Seasonal changes of CO2, CH4 and N2O fluxes in different types of alpine grassland in the Qinghai-Tibetan Plateau of China. Soil Biology and Biochemistry 80: 306–314.
41. Petrakis S, Seyfferth A, Kan J, Inamdar S, Vargas R (2017) Influence of experimental extreme water pulses on greenhouse gas emissions from soils. Biogeochemistry 133: 147–164.
42. Liu C, Holst J, Brüggemann N, Butterbach-Bahl K, Yao Z, Han S, et al. (2008) Effects of irrigation on nitrous oxide, methane and carbon dioxide fluxes in an Inner Mongolian steppe. Advances in Atmospheric Sciences 25: 748–756.
43. Fang H, Cheng S, Yu G, Cooch J, Wang Y, Xu M, et al. (2014) Low-level nitrogen deposition significantly inhibits methane uptake from an alpine meadow soil on the Qinghai-Tibetan Plateau. Geoderma 213: 444–452.
44. Li X, Cheng S, Fang H, Yu G, Dang X, Xu MJ, et al. (2015) The contrasting effects of deposited NH4+ and NO3-on soil CO2, CH4 and N2O fluxes in a subtropical plantation, southern China. Ecological Engineering 85: 317–327.
45. Chen W, Zheng X, Chen Q, Wolf B, Butterbach-Bahl K, Brueggemann N, et al. (2013) Effects of increasing precipitation and nitrogen deposition on CH4 and N2O fluxes and ecosystem respiration in a degraded steppe in Inner Mongolia, China. Geoderma 192: 335–340.
46. Carlsen HN, Joergensen L, Degn H (1991) Inhibition by ammonia of methane utilization in Methylococcus capsulatus (Bath). Applied Microbiology and Biotechnology 35: 124–127.
47. Schnell S, King GM (1994) Mechanistic analysis of ammonium inhibition of atmospheric methane consumption in forest soils. Applied and Environmental Microbiology 60: 3514–3521. 16349403
48. Saari A, Rinnan R, Martikainen PJ (2004) Methane oxidation in boreal forest soils: Kinetics and sensitivity to pH and ammonium. Soil Biology and Biochemistry 36: 1037–1046.
49. Bodelier PLE, Laanbroek HJ (2004) Nitrogen as a regulatory factor of methane oxidation in soils and sediments. FEMS Microbiology Ecology 47: 265–277. doi: 10.1016/S0168-6496(03)00304-0 19712315
50. Tian D, Niu S, Pan Q, Ren T, Chen S, Bai YF, et al. (2015) Nonlinear responses of ecosystem carbon fluxes and water use efficiency to nitrogen addition in Inner Mongolia grassland. Functional Ecology 30.
51. Wolf B, Zheng X, Brüggemann N, Chen W, Dannenmann M, Han XG, et al. (2010) Grazing-induced reduction of natural nitrous oxide release from continental steppe. Nature 464: 881–884. doi: 10.1038/nature08931 20376147
52. Conrad R (2007) Microbial Ecology of Methanogens and Methanotrophs. Advances in Agronomy: Academic Press. pp. 1–63.
53. St Pierre KA, Danielsen BK, Hermesdorf L, D'Imperio L, Iversen LL, Elberling B (2019) Drivers of net methane uptake across Greenlandic dry heath tundra landscapes. Soil Biology and Biochemistry 138: 107605.
54. Liebig MA, Gross JR, Kronberg SL, Phillips RL, Hanson JD (2010) Grazing management contributions to net global warming potential: A long-term evaluation in the northern great plains. Journal of Environmental Quality 39: 799–809. doi: 10.2134/jeq2009.0272 20400576
55. Saggar S, Hedley CB, Giltrap DL, Lambie SM (2007) Measured and modelled estimates of nitrous oxide emission and methane consumption from a sheep-grazed pasture. Agriculture, Ecosystems and Environment 122: 357–365.
56. Whiting GJ, Chanton JP (1993) Primary production control of methane emission from wetlands. Nature 364: 794–795.
Článek vyšel v časopise
PLOS One
2019 Číslo 12
- Řekněte mi, co jíte, a já vám řeknu, jaká je úroveň vaší kognice
- Původní český test ALBA pro odhalení mírných kognitivních deficitů získal certifikaci MZ ČR
- Duševní onemocnění se u teenagerů přenáší v rámci školní třídy
- Od kostiček k lidské kůži. Jak si postavit 3D biotiskárnu?
- Prof. Libor Grubhoffer: Nová varianta mpox je příležitostí pokročit v poznání původce varioly
Nejčtenější v tomto čísle
- Methylsulfonylmethane increases osteogenesis and regulates the mineralization of the matrix by transglutaminase 2 in SHED cells
- Oregano powder reduces Streptococcus and increases SCFA concentration in a mixed bacterial culture assay
- The characteristic of patulous eustachian tube patients diagnosed by the JOS diagnostic criteria
- Parametric CAD modeling for open source scientific hardware: Comparing OpenSCAD and FreeCAD Python scripts