Impact of biochar and lignite‐based amendments on microbial communities and greenhouse gas emissions from agricultural soil

Understanding the responses of the microbial community and greenhouse gas (GHG) emissions to the incorporation of different organic amendments is essential for their proper utilization. In this study, laboratory‐incubated microcosm experiments were conducted to investigate the short‐term effects of pine‐wood biochar and lignite‐based amendment on the microbial communities and GHG emissions from agricultural soil. Soils amended at five different application rates were incubated for 19 d under the conditions of 60% water‐filled pore space and 25 °C. Microbial biomass in the amended soil after incubation was measured by the solid colony counting method, and the soil microbial diversity was assayed using a Biolog EcoPlate. The biochar and lignite‐based amendment had distinct effects on the soil microbial communities and GHG emissions. The microbial community growth and utilization of C sources were improved by the biochar but restrained by the lignite‐based amendment in most cases. The biochar and lignite‐based amendment had a minor impact on methane emissions. Carbon dioxide emissions were promoted by the biochar and inhibited by the lignite‐based amendment during the short‐term incubation period. Nitrous oxide emissions decreased with the application rate of biochar but increased with the rate of lignite‐based amendment. The addition of biochar at a rate of 3–4% and lignite‐based amendment at a rate of <1% has the potential to improve soil quality. Salt leaching is required to avoid accumulation when the biochar and lignite‐based amendments are applied. The findings can provide a reference for the application of biochar and lignite‐based amendment in silt loam soil.


INTRODUCTION
Agricultural development in arid and semiarid regions is threatened by soil salinization, fertility limitations, and water shortages (Sakadevan & Nguyen, 2010). In order to obtain high crop yields and meet the projected increase in human food demand, excessive mineral fertilizers are frequently applied in these regions. However, excessive application of mineral fertilizers may contribute to a series of environmental problems such as non-point-source pollution, soil degradation, desertification, growth of soil-borne greenhouse gas (GHG) emissions, and groundwater contamination in the shallow water zone. To maintain soil health and sustainable agricultural development, diverse organic amendments made from agricultural and industrial byproducts are used to improve the soil fertility and minimize land degradation. The addition of these organic amendments is regarded as an effective method to improve soil structure and fertility, soil C content, microbial biomass and activity, and plant growth and yield. The application of soil amendments is also considered an approach to sequester C and mitigate climate change (Jeffery et al., 2011;Lehmann et al., 2011). However, organic amendments addition may also have disadvantages, such as GHG emissions and nutrient eutrophication resulting from overapplication and misuse (Thangarajan et al., 2013). Therefore, understanding the response of the microbial community and GHG emissions to different organic amendments is essential for the proper utilization of diverse soil amendments in farmland to improve agricultural production while minimizing the negative effects on the environment. A wide array of organic soil amendments with varying levels of processing and characterization are available in different regions. The physicochemical properties and functions of organic amendments differ as a consequence of several factors such as the feedstocks and production conditions. The potential feedstock of organic amendments may originate from various agricultural (e.g., wood, crop residue, and manure) and industrial (municipal solid waste compost, sewage sludge, fly ash, steel mill slag, and lignite) byproducts (Thangarajan et al., 2013). Biochar is a C-rich amendment produced by pyrolysis of organic biomass under the complete absence (pyrolysis) or partial absence (gasification) of oxygen at temperatures ranging from 300 to 1,000˚C (Saifullah et al., 2018). The physicochemical properties of biochar differ as a consequence of the feedstock composition, pyrolysis temperature, and pyrolysis duration (Mcbeath et al., 2014). Biochar is generally characterized by a large surface area, abundant micropores, and a high cation exchange capacity (Lee et al., 2013). Low-cost brown coal (lignite) is another organic matrix that may be suitable as a potential soil amendment because it is a rich source of humic acid (Tran et al., 2015). Lignitebased amendments have a complex intraparticle pore structure with numerous micropores (0.4-1.2 nm) contributing to a high surface area, and thus a large number of active sites (Tran et al., 2015). As lignite is in the early stages of coalification, the properties of lignite-based amendments can vary widely depending on their original source material and environmental and geological factors. Although biochar and lignite-based amendments have different original feedstocks and processes, they have some common properties, such as a complex interpore structure, large surface area, and C-rich structure. A large

Core Ideas
• Biochar and lignite-based amendment have distinct effects on soil microbial communities. • Biochar and lignite-based amendment have limited impact on CH 4 emissions. • CO 2 emissions are promoted by the biochar but inhibited by the liginite-based amendment. • N 2 O emissions decrease with the biochar but increase with the lignite-based amendment.
amount of research related to the influence of biochar on soil microbiomes and GHGs has been conducted (Amini et al., 2015;Beesley et al., 2011;Jeffery et al., 2011;Lehmann et al., 2011), but a very limited amount of research on lignite-based amendments. Therefore, it is critical to compare their different impacts and active mechanisms on soil microorganism and GHG emissions to properly use these amendments. Different organic amendments have different effects on soil physicochemical properties and hydrological functions (Beesley et al., 2011;Brassard et al., 2016;Saifullah et al., 2018). Blanco-Canqui (2017) reviewed the impacts of biochar on soil properties and discussed the factors determining the performance of biochar. Biochar addition may improve soil aggregation and increase the proportion of water-stable aggregates, alter the saturated hydraulic conductivity, increase the water retention capacity, moderate the soil temperature, and reduce abrupt fluctuations in soil temperature (Blanco-Canqui, 2017;Głąb et al., 2016). However, these performances vary with the feedstock type, pyrolysis temperature, particle size, and time after biochar application (Baiamonte et al., 2019;Burrell et al., 2016;Jeffery et al., 2015;Ojeda et al., 2015). Soil physicochemical properties are modified in the presence of lignite (Clouard et al., 2014). The porosity of soil with lignite is observed higher than that of soil without lignite. Most lignite has a complex interpore structure, large surface area, C-rich structure, and organic matrix contributing to increase in the organic C content, C/N ratio, and cation exchange capacity (Clouard et al., 2014;Detman et al., 2018;Tran et al., 2015). In addition, primary organic constituents such as cellulose and the more recalcitrant lignin, humins, or humic acids are commonly present in lignite. Thus, the addition of a lignite amendment was found to promote soil humic acids (Kwiatkowska et al., 2008). Therefore, the influence of different organic amendments on soil properties and hydrological functions varies with the original feedstock and generation processes.
Organic amendments are most frequently used to rebuild the soil organic matter content, provide essential nutrients (such as N, P, and K), and reestablish microbial populations (Larney & Angers, 2012). The application of biochar to soils changes the soil physicochemical properties and stimulates the activities of soil microorganisms (Lehmann et al., 2011). Biochar amendment changes microbial habitats, directly or indirectly affects microbial metabolic activities, and modifies the soil microbial community in terms of abundance and structure (Palansooriya et al., 2019). The addition of biochar to soil often results in positive effects on the abundance of fungi (Warnock et al., 2007), but negative effects have also been observed in biochar-amended soil . Biochar-amended soils have also shown significant changes in the community composition and diversity of fungal, bacterial, and archaeal populations. Xu et al. (2014) found that biochar application significantly increases the diversity of soil bacteria and changes the relative abundance of some microbes related with carbon and nitrogen turnover. Q.  reported that biochar amendment increases bacterial diversity but decreases soil microbial biomass. Senbayram et al. (2019) found that biochar addition has a more distinct influence on the bacterial community composition in sandy soil than in clay soil. However, no discernible differences in the bacterial community structure or even lower diversity of archaea and fungi were found in the biochar-amended soil (Anderson et al., 2014). Therefore, the effect of biochar on microbial performance is determined by a variety of physical and chemical properties of biochar, as it may provide different habitats for microorganisms (Palansooriya et al., 2019). In contrast, research related to the influence of lignite-based amendments on microbial communities is lacking and not well understood. The incorporation of lignite-based amendments in soil may enhance microbial activity by helping to increase in soil moisture content and providing additional organic C as a substrate for microbiomes. Clouard et al. (2014) reported that the microbiological diversity in lignite-rich soil is lower than that in soil without lignite. However, Rumpel et al. (2001) reported that the presence of lignite does not have a significant impact on the microbial biomass in lignite-containing mine soils. Tran et al. (2015) found that lignite amendment has limited impacts on soil microbial communities, and lignite together with N fertilizer does not significantly stimulate microbial activities.
Agricultural soil-borne GHG emissions are primarily released through plant and microbial processes and affected by soil physical, chemical, and biological properties. Biochar and its storage in soils have been heralded as a solution to mitigate GHG emissions by sequestering C and simultaneously providing environmental and agricultural benefits (Zhang et al., 2019). Karhu et al. (2011) reported that biochar addition to agricultural soil increases CH 4 uptake. Case et al. (2015) found that biochar suppresses N 2 O emissions while maintaining N availability in sandy loam soil. Xu et al. (2014) also found that biochar reduces N 2 O emissions by stimulating both nitrification and denitrification. Woolf et al. (2010) estimated that biochar can sustainably offset 7-12% of anthropogenic GHG emissions on an annual basis without endangering food security or habitat or soil conservation. They clarified that the annual net emissions of CO 2 , CH 4 , and N 2 O can be reduced by a maximum of 1.8 Pg CO 2 -C equivalent (CO 2 -Ce), and total net emissions can be reduced over the course of a century by 130 Pg CO 2 -Ce. In contrast, Verhoeven and Six (2014) debated whether biochar mitigated field-scale N 2 O emissions in a northern California vineyard. Senbayram et al. (2019) reported that the application of alkaline biochar to acidic soils can potentially increase N 2 O and CO 2 emissions. There are few reports on GHG emissions from soil reclaimed by lignite-based amendments. Tran et al. (2015) found that CO 2 emissions from soils amended with lignite are inhibited in the short term.
Although a variety of studies have been conducted to investigate the effects of organic amendments on soil functions, the mechanisms of the effects on soil microbial activities and GHG emissions are still not well understood, in particular for the lignite-based amendments. The performance of amendments also varies with their feedstock, processing, and application rate, as well as with the soil type and in situ conditions. The objective of this study was to investigate the impacts of biochar and a lignite-based amendment on microbial communities and GHG emissions from agricultural soil through laboratory-incubated microcosm experiments. We hypothesize that the lignite-based amendment and biochar may have distinct effects on soil microbiomes and GHG emissions even though they have a variety of common properties. The specific purposes were (a) to investigate the effects of biochar and a lignite-based amendment on soil properties; (b) to understand the impact of the biochar and a lignite-based amendment on soil microbial communities and the diversity of microbial community functions; (c) to explore the influence of the biochar and a lignite-based amendment on GHG emissions from soil; and (d) to analyze the relationships between soil microbial activities and GHG emissions from soil with different amendments.

Physicochemical properties of soil, biochar, and lignite-based amendment
The soil used in the incubation experiment was sampled from a maize field in the Hetao Irrigation District located in the upper reaches of the Yellow River, China. Research on the effects of irrigation and fertilization regimes on grain yield, water, N productivity, and soil-borne GHG emissions from a mulched cultivated maize field has been conducted in the sampling site (Li, Xiong, Cui, et al., 2020;. Soil was sampled from the experimental plots at The tested soil is classified as anthropogenic-alluvial soil according to Chinese Soil Classification and Terminology (Shi et al., 2010). Texture analysis showed that the soil was silt loam according to the U.S. soil texture triangle. Two different amendments (i.e., biochar and lignite-based amendment) were separately added to the soil. The soil organic matter content was approximately 22.2 g kg −1 prior to the addition of amendments. The biochar was created from pine wood. The chips of pine wood were placed in ceramic crucibles, each covered with a fitting lid, and pyrolyzed in a muffle furnace under oxygen-limited conditions. The pyrolysis temperature was raised to 500˚C at a rate of 20˚C min −1 , and the temperature was then held at 500˚C for 4 h. The lignite-based amendment was produced by the Apaxfon Biological Science and Technologies Company in China. Both the biochar and lignite-based amendment were air dried and passed through 1-mm sieves. The detailed physicochemical properties of the soil, biochar, and lignite-based amendment are presented in Table 1.

Incubation experiments
Incubation experiments were conducted to examine the responses of the microbial community and GHG emissions to the addition of the biochar and lignite-based amendment, respectively. The biochar and lignite-based amendment were separately and uniformly mixed with soil at weight proportions of 1, 2, 3, 4, and 5%, respectively. The addition quantities of these treatments were approximately equivalent to the application rates of 14, 28, 42, 56, and 70 t ha −1 in the field, respectively. Soil without any amendment was considered the control. Soil with different quantities of amendment was packed into a flask with a volume of 650 ml. Each flask was filled with 210 g of soil mixture, and the packing bulk density was 1.4 g cm −3 , which was consistent with the bulk density in the field. Three replicates were conducted for each treatment. Deionized water was uniformly sprayed on the soil surface using a syringe to increase the soil water content to 40% water-filled pore space (WFPS; i.e., degree of saturation). The WFPS was calculated using the following equation: where ω is the mass wetness (g −1 ), ρ b is the soil bulk density (g cm −3 ), f is the soil porosity, ρ s is the density of solids (2.65 g cm −3 ), and ρ w is the density of water (g cm −3 ). The flasks were sealed with Parafilm and pre-incubated in an incubator at 28˚C and a relative humidity of 50% for 7 d. Subsequent to the primary incubation, urea (46% N) fertilizer was dissolved in deionized water and uniformly applied to the soil surface at a rate of 0.32 g N kg −1 . This application was equal to the application of 250 kg ha −1 in the field. The final soil water content was controlled at 60% WFPS. All the flasks were incubated at 25˚C without cover except for the GHG sampling periods. The soil water content was controlled at 60% WFPS by adding deionized water through a syringe every day.
Gases were sampled using a polypropylene syringe and immediately injected into a gas container. The flask was closed during the sampling period. The concentrations of GHGs were examined within 48 h using a GC-2014 gas chromatograph (Shimadzu Scientific Instruments) equipped with an autosampler, thermal conductivity, flame ionization detector (FID), and electron capture detector (ECD). The oven temperature for gas chromatography was controlled at 50˚C. Nitrous oxide was detected with a 63 Ni ECD at 300˚C. Methane and CO 2 were detected by a FID at 200˚C. The fluxes of CO 2 , N 2 O, and CH 4 were calculated by the slope of the linear regression between the GHG concentration and time, which is expressed as follows: where F is the emission flux of the gas (mg m −2 h −1 ), d c /d t is the variation ratio of the measured gas, M is the molar mass of the gas (g mol −1 ), T is the incubation temperature (K), V 0 is the volume of measured gas under the standard condition (m 3 ), and V is the gas space volume of the flask (m 3 ). The cumulative GHG emissions during the incubation were estimated by the following equations: wherēis the average emission flux of the incubation period (mg m −2 h −1 ), d is the days of incubation, F i is the measured gas emission flux for the ith sample (mg m −2 h −1 ), and d i is the number of days between the adjacent sampling events.

Soil biochemistry assay
The soil was collected to examine the biochemical properties after the termination of incubation. The soil electrical con-ductivity (EC) and pH were determined on a 1:5 saturated soil extract. The NH 4 + -N and NO 3 --N contents of the soil were determined from 2 mol L −1 KCl extracts (soil to water ratio of 1:5) and measured using a continuous flow analyzer (AutoAnalyzer 3 HR, Seal Analytical). Soil organic matter was measured using the potassium dichromate volumetricexternal heating method. The available K was examined using the ammonium acetate extraction-flame photometry, and the available P was measured using the sodium bicarbonate extraction-Mo-Sb colorimetry method (Pansu & Gautheyrou, 2006).
The soil microbial biomass was measured using the solid colony counting method. Ten grams of soil sample was suspended in 90 ml of sterile deionized water. Suspensions were shaken for 30 min in a vibrator and diluted with variable dilution factor according to the different types of microorganisms. Each dilution factor was determined on the basis of pre-incubation experiment. The diluted liquid was inoculated to different media by an automatic spiral plating system. The media were incubated 4-5 d in the dark at 28˚C, and then the colonies were observed. Fungi, bacteria, and actinomycetes were incubated in beef peptone agar medium, Martin's medium, and modified Gao's No. 1 medium, respectively. Nitrifying bacteria, NH 3 -oxidizing bacteria, and denitrifying bacteria were selectively incubated in specific media for their physiological groups (Li et al., 2005).
The microbial community functional diversity of soil was assessed using a Biolog EcoPlate with 31 C sources (Garland, 1997). Soil samples were incubated for 24 h at 25˚C, and then 10 g of incubated soil was mixed with 90 ml of 0.85% (m/v) sterile NaCl solution. The mixture was shaken at 22˚C for 1 h at 200 rounds min −1 . After 30 min of settling, 1 ml of supernatant was diluted in 9 ml of 0.85% sterile NaCl solution at 10-fold dilutions until a final 1:1000 dilution was reached (Feigl et al., 2017). Subsequently, 150 μl of suspension was added to the microplate wells and then incubated for 240 h at 25˚C in the dark. The absorbance value of 590 nm was examined every 24 h using the Biolog EcoPlate system (Biolog). All treatments had three replicates in each Biolog EcoPlate.
The microbial community functional diversity is expressed as the average well color development (AWCD). The calculation of AWCD is expressed as where C i is the optical density value at 590 nm for each well, R is the optical density value of the control group, and n is the number of C sources for the Biolog EcoPlate (31). Substrate richness (S) was calculated by counting the total number of C substrates oxidized by individual treatments on the Biolog EcoPlate. Diversity parameters, namely the Shan- where is the ratio of the absorbance value of each well to the sum of the absorbance value of all wells, and n i (= C i − R) is the relative absorbance value of the ith well.

Data analysis
The daily GHG emissions, cumulative emissions, soil organic matter, soil nutrients, and diversity values were compared in different treatments by one-way ANOVA using SPSS package 20.0 (SPSS). The assumptions on normality of residuals and homogeneity of variance were checked firstly. Statistical differences among the treatments were analyzed by Duncan's test at a significance level of P = .05. The AWCD data based on substrate utilization were analyzed using principal component analysis (PCA, Canoco 5.0).

Soil physical and chemical properties
The soil pH was 6.17 prior to incubation and reached 6.36 after incubation for the control. The soil pH nonlinearly increased with the addition of biochar and the lignite amendment ( Figure 1). The soil pH of the lignite-amended soil was slightly higher (average of 0.17 units) than that of the corresponding biochar-amended soil. The soil EC linearly increased with the application rate of biochar and the lignite-based amendment, but the magnitude (or slope) was evidently different. The increase in the EC of the soil amended with the lignite-based amendment was clearly larger than that of the biochar-amended soil. The soil EC reached 1,020.0 μS cm −1 and 2,396.7 μS cm −1 for the treatments with 5% biochar and lignite-based amendment, respectively. In contrast, the soil EC was 765.7 μS cm −1 for the control.
The soil organic matter also significantly (P < .05) increased with the addition of different amendments (Figure 2). The content of organic matter in the soil with lignitebased amendment was clearly higher than that with biochar after the incubation as organic matter content in the original lignite-based amendment (976.0 g kg −1 ) was much higher than that in the biochar (675.7 g kg −1 ). The soil organic matter contents reached 53.48 and 95.86 g kg −1 for the 5% biochar  and lignite-amended soils, respectively. In contrast, the soil organic matter content was 22.2 g kg −1 in the nontreated soil. The total C, NO 3 --N, NH 4 + -N, and available P and K in the soil after incubation with different treatments are presented in Table 2. The soil total C significantly (P < .05) increased with the addition of both the biochar and lignitebased amendment. The highest total C contents were 4.86 and 4.41% for the 5% biochar and lignite-based amendment treatments, respectively. The total N in the incubated soil was not significantly (P > .05) different for the different biochar treatments. The total N in the soils amended with different amounts of biochar varied from 0.18 to 0.20%, which was similar to that in the control (0.19%). In contrast, the soil total N significantly (P < .05) increased from 0.22 to 0.31% with the addition of the lignite-based amendment. The soil NO 3 --N content significantly (P < .05) increased as the application rate of biochar and lignite-based amendment increased. The increase in the soil NO 3 --N content of the lignite-based amendment treatments was higher than that of the biochar treated one. Although the soil NH 4 + -N content in the biochar-treated soils was less than that in the control, the difference was not significant (P > .05). The addition of the lignite-based amendment had no significant effect (P > .05) on soil NH 4 + -N when the application rate was lower than 3%.
However, the soil NH 4 + -N content significantly increased when the application rate was higher than 4%. Soil available P generally increased with the application of biochar. However, the difference was not significant (P > .05) among different biochar treatments when the application rate was lower than 4%. Soil available P significantly (P < .05) increased with the addition of the lignite-based amendments. The available P in the soil amended with 5% lignite-based amendment was approximately fivefold higher than that in the control.
The addition of both biochar and lignite-based amendments increased the soil available K. However, statistical analysis indicated that the increase in soil available K in different biochar treatments was not significantly (P > .05) different from that in the control. A significant increase in soil available K was detected with the addition of lignite-based amendments. Soil available K increased from 1,005.43 mg kg −1 with the 1% lignite-based amendment to 2,058.42 mg kg −1 (5% lignite-based amendment). Table 3 presents the categories and number of microorganisms in the soil after incubation with different additions of biochar and lignite-based amendment. The number of fungi in the biochar-amended soils generally increased with the increase in the biochar application rate, but decreased when the rate was beyond 4%. The largest fungi population (2.43 × 10 4 colony-forming units [CFU] g −1 ) was obtained in the 3% biochar treatment. In contrast, the number of fungi cultures decreased with the addition of lignite-based amendment (except for in the 1% treatment), but the difference was not significant (P > .05). The number of fungi in 1% lignitebased amendment increased by 40.6%, whereas the number of fungi in the treatments with more than 1% lignite-based amendment was reduced on average by 54.1% compared with that of the control.

Microbial communities in soils amended with the biochar and lignite-based amendment
T A B L E 2 Soil total C, diverse forms of N, and available P and K after incubation for different amendment treatments (mean ± standard deviation) The number of actinomycetes nonmonotonically varied with the application rate of biochar. The addition of biochar with the ratio of 3 and 4% promoted the growth of actinomycetes. The number of actinomycetes in the other biochar treatments was less than that in the control. The maximum number of actinomycetes (9.09 × 10 5 CFU g −1 ) was obtained in the 3% biochar treatment. The addition of the lignite-based amendment generally reduced the number of actinomycetes in the soil. The greater the addition of lignite-based amendments, the less the actinomycetes cultures grew. The number of actinomycetes cultures in the 5% lignite-based amendment treatment was reduced by 36% compared with that of the control.

Treatments
The population of bacterial cultures also nonmonotonically increased with the increase in the biochar application rate. The bacteria populations in the 1 and 2% biochar treatments were lower than those in the control (1.34 × 10 8 CFU g −1 ), and those in the 3-5% biochar treatments were significantly (P < .05) higher than that in the control. The largest bacteria population (1.97 × 10 8 CFU g −1 ) was obtained in the 3% biochar treatment. The addition of the lignite-based amendment decreased the bacteria population for most treatments (except for the 2% treatment). The greater the addition of lignite-based amendment (>2%), the lower the bacteria population obtained in the soils. Soil amended with the 5% lignitebased amendment had only 37% of the population of bacteria compared with that in the control.
The number of nitrifying bacteria increased nonmonotonically with the application rate of biochar. The number of nitrifying bacteria increased by 15.1-82.2% in different biochar treatments compared with that in the control. The largest number of nitrifying bacteria was obtained in the 5% biochar treatment with 1.95 × 10 7 CFU g −1 . In contrast, the addition of the lignite-based amendment inhibited the growth of nitrifying bacteria. The number of nitrifying bacteria in all the lignitebased amendment treatments was less than that in the control. The lowest number of nitrifying bacteria was obtained in the treatment with 3% lignite-based amendment, and the highest number of nitrifying bacteria was obtained in the 1% lignitebased amendment.
The number of denitrifying bacteria was three orders of magnitude less than the number of nitrifying bacteria in different treatments. Denitrifying bacteria increased with the addition of biochar. The largest number of denitrifying bacteria was obtained in the 2% biochar treatment. The addition of lignite-based amendment had no significant (P > .05) effects on the denitrifying bacteria. The mean number of denitrifying bacteria was 5.98 × 10 4 CFU g −1 in the soils amended with the lignite-based amendment.
Biochar addition significantly (P < .05) increased the number of soil NH 3 -oxidizing bacteria. Soil treated with 4% biochar had the most NH 3 -oxidizing bacteria. In contrast, the presence of the lignite-based amendment had limited impact T A B L E 3 The category and number of microorganisms in soil after incubation for different biochar and lignite-based amendment treatments (mean ± standard deviation) on the soil NH 3 -oxidizing bacteria (except for the 1% lignitebased amendments treatment). The number of NH 3 -oxidizing bacteria was 1.65 × 10 6 CFU g −1 in the 1% lignite-based treatment, and the mean number of NH 3 -oxidizing bacteria was 1.26 × 10 6 CFU g −1 in the other lignite-based treatments. The community-level physiological profile was characterized by the AWCD and principal components (PCs). Figure 3 illustrates the AWCD in the biochar and lignite-based amendment treatments during the incubation period. The AWCD increased as the incubation period increased for the biochar treatments. The 3 and 4% biochar treatments consistently exhibited higher AWCD than the control. The 4% biochar treatment had the highest AWCD at all sampling periods. The other biochar treatments had a higher AWCD than the control after 168 h of incubation, but a lower AWCD than the control after 192 h of incubation. In contrast, the AWCD increased with the incubation period for the 1 and 2% lignite-based treatments but did not substantially change for the 3-5% treatments. After 120 h of incubation, all the lignite-based amendment treatments had a lower AWCD than the control. The 1% lignite-based amendment treatment consistently exhibited the highest AWCD at all sampling periods. All lignite-based amendment treatments had a lower AWCD than the biochar treatments after 240 h of incubation at a temperature of 25˚C. This indicated that the addition of lignite-based amendment inhibited the activity of soil microbial communities.
Principal component analysis of 31 C sources revealed the different patterns of potential C utilization and different microbial communities. Five PCs with eigenvalues >1 were identified. The percentages of PC1, PC2, PC3, PC4, and PC5 were 48.04%, 22.61%, 12.33%, 7.26%, and 5.31%, respectively. The cumulative contribution of the first and second PCs was 70.65% (Table 4), so the first two PCs were used to represent the microbial community functional diversity in different treatments. The PCA showed distinct differences among the biochar and lignite-based treatment (Figure 4). Biochar addition increased the utilization of amides, amino acids, carbohydrates, and carboxylic acids (Figures 4a  and 4b). In general, treatments with the same amendment clustered together except for 2% biochar treatment (Figure 4a). Biochar and lignite-based amendment treatments were widely separated from each other on the PC1 axis. Most of the biochar treatments and the control were positively correlated with all the loading variables on PC1, whereas all the lignite-based

F I G U R E 3 Average well color development (AWCD) varied with incubation duration in different biochar (BC) and lignite-based (LB) amendment treatments
amendment treatments were negatively correlated with all the loading variables on PC1. Both PC1 and PC2 separated the treatments during the incubation period, but the degree of separation was different. Distribution along PC2 did not reveal distinct patterns during the incubation period. This indicated that the differences in the metabolism of C sources among biochar and lignite-based amendment treatments were mainly reflected in PC1. The addition of biochar significantly increased the richness when the addition proportion of biochar was 4%, whereas there was no significant effect on the richness for the other biochar treatments (Table 5). The richness decreased with the increase in lignite-based amendment. The richness in the 5% lignite-based amendment decreased by 9.67 compared with that in the control.
The Shannon index quadratically (R 2 = .94) increased with the application rate of biochar. The Shannon index in the 3 and 4% biochar treatments was higher than that in the control. The highest Shannon index was obtained in the 4% biochar treatment ( Table 5). The Shannon index decreased linearly (R 2 = .97) with the increase in lignite-based amendment. The Shannon index in the 5% lignite-based amendment treatment decreased by 32.3% compared with that in the control. There was no significant difference (P > .05) in the Shannon evenness among the biochar treatments. The Shannon evenness increased quadratically (R 2 = .78) with lignite-based amendment addition. The largest value of Shannon evenness was obtained in the 3% lignite-based amendment treatment. No significant difference in the Simpson index was found in both the biochar and lignite-based amendment treatments. Only the addition of 3 and 4% biochar significantly increased the McIntosh index, and the other biochar treatments significantly decreased the value of McIntosh index. The largest McIntosh index was obtained in the 4% biochar treatment. The addition of the lignite-based amendment significantly (P < .05) decreased the McIntosh index. A higher McIntosh index was obtained with the lower application rate of the lignite-based amendment. Overall, biochar addition at rates of 3 and 4% enhanced the diversity and evenness indexes of the microbial communities.

Greenhouse gas emissions from amended soils
The temporal dynamic variation in GHG fluxes during the incubation period is presented in Figure 5. The CH 4 fluxes were generally very small for all treatments. A peak in the CH 4 flux occurred after 2 d of incubation for all treatments. The peak values were positive for the control and most biochar treatments. The largest peak (0.37 mg m −2 h −1 ) was obtained in the 4% biochar treatment. In contrast, the peak values of CH 4 flux were negative and of much smaller magnitude for most of the lignite-based amendment treatments. After 3 d of incubation, the CH 4 fluxes approached 0 for all treatments.
Biochar addition clearly enhanced the soil CO 2 emissions. The CO 2 fluxes increased with the incubation time, approached the peak value after 5 d of incubation, and then decreased gradually to steady levels after 15 d of incubation for all the biochar treatments. The peak values of CO 2 fluxes varied from 253 to 316 mg m −2 h −1 for different biochar treatments. In contrast, the addition of the lignite-based amendment inhibited the soil CO 2 emissions compared with the control. The CO 2 fluxes decreased with the increase in the lignite-based amendment. The peaks of CO 2 fluxes occurred after 4-5 d of incubation for all lignite-based amendment treatments. The peaks of CO 2 fluxes varied from 170 to 217 mg m −2 h −1 for different lignite-based amendment treatments. T A B L E 5 Diversity and evenness indices of microbial community for soil amended with biochar and lignite-based amendment after incubation (mean ± standard deviation) Lignite-based (1%) 8.33 ± 1.15d 2.52 ± 0.07c 1.20 ± 0.09c 0.90 ± 0.03bc 1.04 ± 0.18f
Biochar hindered the soil N 2 O emissions compared with the control. A peak in N 2 O fluxes occurred after 16 h of incubation in all biochar treatments, and then decreased gradually to the steady levels (4 d after incubation). The peak values of N 2 O fluxes varied in the range of 3.9 to 19.6 mg m −2 h −1 in different biochar treatments. The N 2 O fluxes in the treatments of lignite-based amendment decreased steadily over time and stabilized after 4 d of incubation. The main emission periods of N 2 O occurred in the first 4 d of incubation for different treatments with biochar and lignite-based amendment adding. However, the N 2 O fluxes in the biochar treatments were much smaller than those in the lignite-based amendment treatments in the initial incubation stage (2 d after incubation).
The cumulative GHG emissions during the incubation period are presented in Figure 6. The cumulative CH 4 emissions were approximately 0.01 kg ha −1 in biochar treatments and −0.002 kg ha −1 in lignite-based amendment treatments. The cumulative CO 2 emissions nonlinearly increased with the rate of biochar addition. The maximum cumulative CO 2 emissions were obtained in the 4% biochar treatment, which increased by 60.6% compared with the control. In contrast with biochar treatment, the cumulative CO 2 emissions generally decreased with the addition of lignite-based amendment. The cumulative CO 2 emissions from soil amended with 5% lignite-based amendment decreased by 19.6% compared with the control. In addition, the cumulative CO 2 emissions in biochar treatments were larger than those in the lignite-based amendment treatments. The addition of biochar generally decreased the cumulative N 2 O emissions, but the cumulative N 2 O emissions increased with the application rate of the lignite-based amendment. The maximum cumulative N 2 O emissions reached 1.94 kg ha −1 (5% lignite-based amendments) after 19 d of incubation.

DISCUSSION
This study investigated the effects of biochar and lignitebased amendment addition on soil physicochemical properties, microbial communities, and GHG emissions through laboratory-incubated microcosm experiments. Soil physicochemical properties varied with the addition of biochar and lignite-based amendment. The soil microbial abundance and utilization ability of C sources were improved by the biochar in general but restrained by the lignite-based amendment in most cases. The GHG emission features were distinct between the biochar and lignite-based amendment treatments due to the different influences on soil microbial communities.

Impact of organic amendments on soil properties
A large amount of research shows that soil physicochemical properties are improved with the biochar incorporation (He et al., 2020;Lehmann et al., 2011;McHenry, 2011). Most of these findings have been obtained from pot and field experiments, which are influenced by a variety of factors such as climate, soil, crop, and field management (Gul et al., 2015;Laird et al., 2010;Yao et al., 2017). In the present study, we investigated soil physicochemical properties response to the biochar and lignite-based amendment through laboratory-incubated microcosm experiments. The addition of the biochar and lignite-based amendment resulted in the increase in soil pH, EC, organic matter, and available nutrients. However, the increase magnitude in lignite-based amended soils was clearly larger than that in the biochar-amended ones due to its original high background.   Soil C/N ratio increased with biochar addition but decreased with the application of lignite-based amendment. The addition of biochar with a high C/N ratio resulted in a high C/N ratio in the biochar-amended soil, which is consistent with the findings of Case et al. (2012) and Hu et al. (2014). The decreased C/N ratio in soil with lignite-based amendment addition is different from the findings of Zhong et al. (2010), who found that N-modified lignite with a C/N ratio of 76.2 significantly increased the soil microbial biomass C/N ratio (the original C/N is 8.5). The decrease in C/N Bars with a common letter are not significantly different at the .05 level by Duncan's test ratio in the present study is attributed to the original low C/N ratio that is lower than the tested soil. The difference of biochar and lignite-based amendment in soil C/N ratio is attributed to the original C/N ratio (Table 1). As shown in Gelsomino et al. (2006), the C/N ratio in the amended soil is highly dependent on the original C/N ratio of the amendments. The addition of biochar and lignite-based amendment enhanced soil EC. The increase in EC in lignite-based amended soil was greater than that in biochar amended ones as the higher initial background (Table 1). In addition, the lignite-based amendment had a much higher organic matter content (976.0 g ka −1 ) than the biochar (675.7 g kg −1 ), so the decomposition of organic matter may release ions and increase the soil EC (Abdelhafez et al., 2014). The increase in soil EC with the addition of biochar and lignite-based amendment was also detected in previous field and incubation experimental studies (C. Masto et al., 2013;Zolfi-Bavariani et al., 2016). A high salt content may be harmful to crops and soil microorganisms. In the short-term incubation period (20 d), the EC in biochar-amended soils ranged from 766 to 1,020 μS cm −1 , and varied from 1,036 to 2,396 μS cm −1 in lignite-based amendment treated ones. Ezrin et al. (2010) reported that the threshold of soil EC for crop growth and yield is 1,150 μS cm −1 . Clearly, the EC of the lignitebased amendment treated soil was beyond the threshold of 1,150 μS cm −1 . Although the EC in the biochar amended soil is less than the threshold of 1,150 μS cm −1 , secondary salinization and salt harm to crop growth and microorganisms should be considered for the incorporation of biochar and lignite-based amendment, in particular, for the prolonged application of these amendments.

Effects of the biochar and lignite-based amendments on soil microbial communities
The species abundance and structure of the soil microbial community varied with the application of different organic amendments at different application rates (Tables 3 and 4). The application of biochar at a rate of 3-4% obtained the largest abundance of soil microbial biomass. The addition of biochar promoting the development of microbiomes may occur through providing favorable soil water, nutrients, and aeration environment and increasing organic C and the C/N ratio. Biochar contains abundant macro-and micropores; thus, the addition of biochar increases the total surface area and pore size distribution (Jindo et al., 2014). The improvement of soil surface area and pore structure provide microhabitats for microorganisms (Ameloot, Graber, et al., 2013;Kraychenko et al., 2019), as many soil microorganisms are specialists living in microhabitats that provide resources for their specific metabolic needs (Lehmann et al., 2011). Biochar also contains a small amount of labile organic C and micro-and macronutrients, which are likely to increase microbial abundance (Zavalloni et al., 2011). In addition, biochar addition significantly increased the soil C/N ratio, which could stimulate microorganism growth (Liu et al., 2009). Liang et al. (2017) has reported that the addition of substrate with high C/N ratio can result in a greater priming effect. These direct and indirect factors enhance the microbial abundance after biochar application. On the other aspect, salt and other toxic compounds (such volatile organic compounds, polycyclic aromatic hydrocarbons, and dioxins, and heavy metals, etc.) were added into the soil in companion with biochar application (Zheng et al., 2019), which might have inhibited microbial community growth (Rath et al., 2016). Excessive addition of biochar may inhibit the microbial growth due to the toxicity effect of salt and heavy metals (Shi et al., 2020). In the present study, the numbers of total bacteria decreased when the biochar addition rate was <3%. Ameloot, Graber, et al. (2013) found that a high rate of biochar application (90 t ha −1 ) reduces the number of soil microorganisms.
In contrast, the addition of lignite-based amendment significantly decreased the abundance of the soil microbial community in most cases. Although the soil nutrients increased with the increase in lignite-based amendment addition, soil EC also increased significantly. As many soil microorganisms are negatively affected by salinity owing to water availability restriction as a result of low osmotic potentials in soils and through ion toxicity (Rath & Rousk, 2015). Moreover, a large number of compounds such as phenols, phthalates, polyaromatic hydrocarbons, benzenes, and aliphatic may be applied to the soil (Tran et al., 2015). Some of these components are known to have toxic effects on cellular metabolism (Maharaj et al., 2014). In addition, lignite-based amendment addition generally decreased the soil C/N ratio, which might have constrained the growth of microorganisms. The limited impacts of lignite-based amendments on the microbial abundance are consistent with the findings of Tran et al. (2015).
The abundance of denitrification bacteria was much lower than that of nitrifying and NH 3 -oxidizing bacteria (Table 3). Nitrification is the main N turnover process when the soil WFPS is in the range of 35-60%, whereas denitrification is the main N turnover process when the soil WFPS is >70% (Yoo et al., 2018). The WFPS was controlled at 60% during the experimental period; thus, the abundance of bacteria in the nitrification process was higher than that in the denitrification process.
The Biolog EcoPlate data clearly showed that there was a significant successional shift in the microbial community structure owing to the addition of biochar and lignite-based amendment (Figure 4a). The addition of the biochar and lignite-based amendment may change the substrate use patterns of microorganisms and influence the structure of the soil microbial community (Pietikäinen et al., 2000). In the present study, biochar addition enhanced the ability of soil microorganisms to utilize C sources, whereas the addition of lignite-based amendment inhibited the microbial utilization of soil C sources (Figure 4b). Microbial groups are generally considered to rapidly adapt to changes in soil environmental conditions (Spyrou et al., 2009). However, some research reported that the addition of stable C sources alters the structure of the soil microbial community rather than the total biomass of microbial communities, and a new equilibration may be achieved again after long-term incubation (Anders et al., 2013;Tran et al., 2015). Therefore, the longterm effects of biochar and lignite-based amendment addition on the microbial community structure are desirable.

Influence of organic amendments on greenhouse gas emissions
The addition of amendment altered the soil physicochemical properties and microbial communities, thereby influencing soil-borne GHG emissions. The CH 4 emissions were very small during the incubation period, and soil even acts as a CH 4 sink for some treatments (Figures 5 and 6). Strictly anaerobic condition is required for methanogenesis through different pathways. As soil moisture was maintained at 60% WFPS for all treatments during incubation, the soils were not strictly anaerobic; thus, the methanogenesis was hindered. In addition, the optimum methanogenesis temperature is between 30 and 35˚C (James et al., 1996). After incubation at 25˚C, the activities of methanogens decreased and CH 4 production was reduced.
The addition of biochar significantly increased the cumulative CO 2 emissions at different application rates, whereas lignite-based amendment addition generally decreased the cumulative CO 2 emissions ( Figure 6). The increase in soil respiration is usually attributed to the increased concentration of labile organic C with the addition of amendment and the enhanced soil microbial activities . In contrast, the decrease in soil respiration is often explained by the toxic effect of organic amendments (Zimmerman et al., 2011) and reduction in substrate availability, microbial abundance, and enzymatic activity (Lehmann et al., 2011). Biochar addition increased the microbial abundance (Table 3), altered the microbial community structure, and enhanced the soil C source utilization ability of soil microorganisms (Figure 4). Therefore, biochar addition improved the soil respiration and increased the soil CO 2 emissions. This result is consistent with the findings of Zhou et al. (2017) and Czekała et al. (2016). In contrast with biochar, although organic matter and available nutrients were added to the soil with lignite-based amendment, a large number of salts are also applied to the soil simultaneously ( Figure 1). Both bacterial growth and decomposition are directly inhibited by high salt concentrations, so that soil respiration is constrained (Rath et al., 2019). The addition of lignite-based amendment significantly reduced the microbial abundance in most cases and inhibited the ability of microorganisms to utilize C sources, thereby reducing the soil CO 2 efflux. This is consistent with the findings of Tran et al. (2015), who found that the application of Victorian lignite reduces CO 2 emissions. Schefe et al. (2008) also found that the application of lignite can suppress soil respiration in the short term.
The addition of biochar generally decreased the N 2 O emissions, whereas lignite-based amendment addition increased the N 2 O emissions for most cases (Figure 6). Nitrous oxide emissions decreased in biochar-amended soils and increased in lignite-based amendment soil are consistent with the findings of Bruun et al. (2011) and Sun et al. (2016). Nitrous oxide is a product of microbial metabolism during the process of nitrification and denitrification. The conditions necessary for N 2 O emission in soils include (a) a supply of decomposable organic C, (b) a supply of mineral N, (c) low pH, and (d) the presence of denitrifying organisms (Ameloot, De Neve, et al., 2013). Biochar addition decreased the N 2 O emissions as a result of several possible mechanisms. First, although biochar addition increased the number of nitrification bacteria (Table 3), the NH 4 + -N content decreased with biochar application (Table 2). The low NH 4 + -N content could not provide sufficient substrate for the nitrification process. Second, denitrification rates are usually assumed to decrease with pH increasing (Simek & Cooper, 2002). The increase in soil pH with the addition of biochar decreased the N 2 O emissions (Figures 1 and 6). Third, the porous structure of biochar may increase soil aeration and suppress denitrification (Case et al., 2012). A fourth mechanism might be the absorption of NO 3 --N on the biochar surface leading to the decrease in the N 2 O emissions (Karhu et al., 2011). Although lignite-based amendment addition generally decreased the number of nitrification bacteria ( Table 3). The soil NH 4 + -N content significantly increased, especially in the higher lignite-based amendment treatments which promoted N 2 O emissions ( Figure 6). The increase in N 2 O emissions with lignite-based amendment addition might be caused by the decoupling of either nitrification or denitrification processes because of the significant increase in soil EC (Low et al., 1997). A high salt concentration in soil may inhibit the activity of N 2 O reductase, which results in N 2 O accumulation from denitrification under saline conditions (Menyailo et al., 1997). In addition, biochar addition significantly increased the soil C/N ratio, whereas the addition of lignite-based amendment generally decreased the soil C/N ratio (Table 2). Nitrous oxide emissions are influenced by the soil C/N ratio, since soil N compounds are primary terminal electron acceptors and C acts as an electron donor in the denitrification process (Thangarajan et al., 2013). In general, N 2 O emissions are negatively correlated with the C/N ratio of soil (Huang et al., 2004;Thangarajan et al., 2013). The addition of biochar increased the soil C/N ratio and the incorporation of lignite-based amendment decreased the soil C/N (Table 2). Therefore, N 2 O emission was promoted in lignite-based amendment treated soil but decreased in biochar-amended soil.

CONCLUSIONS
The effects of biochar and lignite-based amendment on soil physicochemical properties, microbial communities, and GHG emissions from agricultural soil were investigated through laboratory-incubated microcosm experiments. The soil pH and EC clearly increased with the addition of biochar and lignite-based amendment. The values of pH and EC in the soil amended with the lignite-based amendment were higher than those in the soil with the same rate of biochar. The incorporation of biochar and lignite-based amendment can improve soil fertility by enhancing the soil organic matter and available nutrients. The C/N ratio in the amended soil is highly dependent on the original C/N ratio of the amendments. The biochar and lignite-based amendment had distinct effects on the soil microbial communities and GHG emissions. The soil microbial abundance and utilization ability of C sources were improved by the biochar but restrained by the lignite-based amendment in most cases. The CH 4 emissions were very small during the incubation period, and soil even acts as a CH 4 sink for some treatments. Carbon dioxide emissions were promoted by biochar and inhibited by the lignite-based amendment during the short-term incubation period. Nitrous oxide emissions decreased with the addition of biochar but increased with the application of lignitebased amendment. As the soil EC significantly increased with the addition of amendments, particularly for the lignite-based amendment. Thus, salt leaching is required to avoid salt accumulation when the biochar and lignite-based amendment are applied in the field. The findings of this research can provide a reference for the application of biochar and lignite-based amendment in silt loam soil. However, this research focused on the individual effects of biochar and lignite-based amendment, the synergistic effects of two organic amendments are still not known. In addition, the long-term effects of the amendments on soil properties, biological processes, and GHG emissions still need to be determined.