Biochar accelerates straw decomposition and reduces greenhouse gas emissions by driving microbial community dynamics

Abstract

Background

See More

The rational utilization of agricultural straw is crucial for improving soil fertility and reducing greenhouse gas emissions (GHGs). The purpose of this study was to investigate how rice (RB) and maize (MB) straw-derived biochar, produced at varying pyrolysis temperatures and application rates, regulated straw decomposition and GHGs by reshaping soil microbial communities and physicochemical properties.

Results

Through 90-day incubation experiments, it was found that biochar produced using low temperature (300 °C) at 2.5–5.0% application rates significantly accelerated straw decomposition by 14.94–36.04% and reduced CH₄ and N₂O emissions by up to 37.84–90.26% and 41.60–91.10%, respectively. Application of biochar produced using low-temperature method enhanced the soil organic matter (9.92–29.26%), pH (1.82–11.32%), and soil enzyme activities (cellulase: 7.84–22.90%, β-glucosidase: 49.92–75.32%), while altering microbial communities, especially increasing copiotrophic bacteria (e.g., Proteobacteria, Ascomycota) in rice grown soils linked to rapid decomposition and reducing Ascomycota dominance in maize soil with altering nutrient dynamics due to higher C/N ratios. Path analysis indicated strong biochar–enzyme–decomposition linkages (normalized coefficient: 0.92), emphasizing microbial community structure as a pivotal mediator. In contrast, biochar produced through high pyrolysis temperatures (mentioning the temperature) diminished effectiveness due to higher structural stability and potential limitations in microbial activity.

Conclusions

Our results indicate that application rates of 2.5–5.0% biochar produced through low temperature can effectively balance straw decomposition and GHGs reduction, offering a sustainable approach for straw management in rice and maize cultivation. These findings provide scientific support for optimizing biochar use in agriculture, contributing to improved soil health and climate change mitigation.

Graphical Abstract

Introduction

China is a major agricultural country, producing 797 million tons and 802 million tons of grain in 2020 and 2021, respectively, with diverse types of straw (http://www.stats.gov.cn/). Furthermore, the rate of agricultural residue generation increased by approximately 4% per year on an average [1]. According to Jin et al. [2], straw is a carbon-rich agricultural waste containing large amounts of nitrogen (N), phosphorus (P), potassium (K), and trace elements, serving as an important nutrient source for crop growth. Straw return to the field is one of the most widely adopted practice for comprehensive utilization of crop residues currently.

Straw return to the field can enhance soil fertility, increase soil organic carbon storage, and improve crop yields [3, 4]. However, traditional direct straw return methods still have many issues in terms of straw resource utilization. For example, in agricultural fields, straw mulching and shallow tillage are commonly used, which often result in slow straw decomposition, insufficient nutrient absorption and utilization by crops from the straw, inadequate benefits from increased soil organic matter, exacerbated pest and disease issues, and reduced crop yields [5, 6]. In addition, after crop straw enters the soil, it undergoes mineralization and humification processes [7]. The action of micro-organisms during the mineralization process releases intricate components in straw—mainly referring to complex biopolymers, such as lignin, cellulose, hemicellulose, and their derived intermediate organic compounds—into the atmosphere in the form of methane (CH₄) and carbon dioxide (CO₂) [2, 8]. Therefore, the most pressing demands for straw to be put back into the field are ways that promote decomposition and lower greenhouse gas emissions (GHGs).

Nowadays, pyrolyzing organic materials from agricultural wastes to create biochar is gaining more and more attention globally [9, 10]. Using biochar as an additive for agricultural soil helps optimize soil structure, maintain soil quality, productivity, increase soil carbon sequestration, and reduce GHGs [11]. The abundant functional groups and porous nature of biochar offer secure anchoring points and nourishment for micro-organisms, thereby boosting the quantity and variety of microbes. The enrichment of microbial communities and the enhancement of microbial activity undoubtedly facilitate the decomposition of organic matter. For example, recent studies have shown that biochar enhanced the abundance of nitrogen-fixing bacteria (e.g., Azotobacter) and phosphorus-solubilizing micro-organisms (e.g., Bacillus) through their nutrient-retaining properties, leading to increased soil ammonium nitrogen and available phosphorus [12]. These micro-organisms accelerate the decomposition of organic matter and nutrient cycling through extracellular enzymes (such as cellulase and urease), with their enzyme activity increasing by 40–80% in soils amended with biochar [13]. The influence of different additions of biochar on dissolved organic carbon (DOM) and total organic carbon (TOC) in compost were diverse, yet the overall trend appeared to be an upward one [14]. Furthermore, by increasing the rate of aeration and decreasing the bulk density of organic material, biochar could lower the production of an anaerobically zone and, consequently, GHGs [15, 16]. Biochar suppresses methanogenic archaea (e.g., Methanobrevibacter) by altering soil redox conditions, reducing methane emissions by 39–50% in paddy soils [17]. At the same time, biochar encourages the activity of denitrifying bacteria with the nirS gene (e.g., Pseudomonas) that transform nitrate into nitrogen, thereby decreasing nitrous oxide (N₂O) emissions by 12–20% in neutral pH soils [13]. These findings highlight the dual roles of biochar in enhancing soil fertility through microbial-driven nutrient cycling while reducing GHGs via targeted modulation of microbial communities.

Numerous studies have documented biochar’s ability to accelerate organic material decomposition [18]. Reported that adding biochar significantly increased the decomposition rate of maize straw. The abundant porous structure of biochar provides habitats for micro-organisms, expanding the contact interface between micro-organisms and straw; its surface functional groups may participate in the initial degradation of complex components, such as lignin and cellulose, synergistically accelerating decomposition with microbial activity [19]. Demonstrated that adding biochar to pig manure treatment systems not only accelerated the decomposition process but also significantly reduced methane and nitrous oxide emissions, achieving rapid nutrient transformation and mitigating environmental impact [20]. Identified that adding biochar can significantly improve the decomposition conditions of municipal sludge; biochar can adsorb toxic substances in the sludge, reduce microbial inhibition, stimulate microbial activity, thereby accelerating the decomposition of organic components, which helps convert sludge into organic fertilizer with ameliorative functions, achieving waste resource utilization.

Based on the above information, we speculate that applying biochar together with straw would accelerate the straw decomposition and lower GHGs from farmland. Nevertheless, the quantitative relationships between straw decomposition rates with biochar’s dosage, and pyrolysis temperature, as well as the influences between individual biochar types on GHGs produced during the straw decomposition, are unclear. The underlying mechanisms including modifications in enzyme activity and composition of microbial communities in the composite crop straw–biochar system should be further explored. Rice and maize are the basic staple crops for approximately half of the world’s population; these two types of straw were chosen as experimental materials in this study. The specific goals were to: (1) examine the impact of the pyrolysis temperature and biochar application rate on straw decomposition and GHGs during the incubation; (2) explore the introduction of various types of biochar altered soil physicochemical properties; and (3) clarify the potential microecological mechanisms underly the influences of biochar addition. The findings of this study could provide the conceptual and technological backing for the scientific validation of straw’s use in agricultural production.

Materials and methods

Materials

Soils were collected from paddy fields in Liuyang, Hunan Province, and maize fields in Taian, Shandong Province, representing typical red paddy soils and yellow–brown maize soils in China. Basic soil properties of the experimental fields were pH 5.28 and 6.57, soil organic carbon 17.55 and 18.40 g/kg, total nitrogen 2.14 and 1.20 g/kg, total phosphorus 1.14 and 0.20 g/kg, and total potassium 15.37 and 10.00 g/kg, respectively. Before processing, the soil was air-dried, sieved (2 mm), and 0.6 kg of soil was placed in each incubation container.

Rice and maize straws were collected from the same sites to ensure consistency between plant residues and soil attributes. These two residues were selected because they are among the most abundant crop straws in China, yet their recycling rate remains low, and direct return often causes decomposition and emission problems. The straws were cut into approximately 5 cm fragments, which had the elemental compositions of 33.54% C and 1.58% N for rice straw, and 43.10% C and 0.87% N for maize straw.

Biochar was prepared from rice and maize straws, separately in a muffle furnace under continuous N₂ flow. The straws were pyrolyzed at 300, 500, and 700 °C with a heating rate of 15 °C/min and holding time of 2 h. The resulting rice–straw biochar (RB) and maize–straw biochar (MB) were stored in airtight bags in the dark. Their basic characteristics, including pH, elemental composition, C/N ratio, and surface area, are summarized in Table S1 (Supplementary file). Detailed measurements of heavy metals, functional groups, and morphology were conducted by standard spectroscopic and microscopic methods (Supplementary Note 1). Baseline measurements of soil enzyme activities and microbial community composition were taken at day 0 before the start of incubation and used as reference values.

Experimental design

For the incubation experiment, rice and maize straws were cut and enclosed in nylon mesh bags (120 mesh, 9 × 7 cm). This method allowed the straws to be in direct contact with the surrounding soil while facilitating subsequent retrieval for analysis. The calculation of straw return amount followed the basis described in Supplementary Note 2. Each bag was buried in a pot (height 199 mm, diameter 136 mm) filled with soil, with straw added at rates equivalent to 14 g/kg for rice and 13 g/kg for maize. The biochars prepared at three different pyrolysis temperatures (300, 500, and 700 °C) were mixed with straw at the rates of 0, 2.5, 5, and 10% (w/w). These levels were chosen to represent low, medium, and high application rates commonly used in field studies, thereby enabling comparison of dose effects. In total, ten treatments with three replicates each were established: control (only straw, CK), 2.5–300, 5–300, 10–300, 2.5–500, 5–500, 10–500, 2.5–700, 5–700, and 10–700. The exact amounts of straw and biochar used in each treatment are listed in Table 1. The incubation was conducted under controlled conditions at ambient temperature (20 ± 1 °C) and soil moisture maintained at approximately 60% of water-holding capacity to simulate typical field conditions while ensuring microbial activity.

Table 1 Amount of biochar and straw added in per treatment

Sampling methods

Soil and straw samples were collected at fixed intervals of 5, 10, 15, 30, 45, 60 and 90 days for all treatments, irrespective of the biochar application rate and pyrolysis temperature. Soil samples were taken by five-point sampling method, which were then split into two halves and kept in a refrigerator at 4 °C and − 20 °C for physicochemical and microbiological analyses, respectively. Rice and maize straw were sampled by retrieving the nylon mesh bags buried in the soil at each designated interval, which contained the straw materials for subsequent washing, drying, and analysis. The straw samples were washed and dried, with stored at 25 °C.

The treatment was placed into a cylindrical compost bucket with an inner diameter of 0.5 m and a height of 0.5 m) sealed for 6 h. The gas collected bag was attached to the three-way valve after 6 h, and the gas-tight locking syringe was used to collect the gas for analysis. The concentration of gas phase samples was determined by gas chromatography (GC).

Analytical methods

Straw sample

A modified version of Van Soest technique [21] was applied to measure the amounts of cellulose, hemicellulose, and lignin in the straw. The methods used for estimating the chemical composition of straw are described in Supplementary Note 3. The weight loss ratio of straw was calculated using the differential weight method, using the following equation:

$${\text{Straw decomposition rate}} = \frac{{\left( {M_{0} - M_{t} } \right)}}{{M_{0} }} \times 100\%$$

wherein M₀ is the original dry weight of the supplied straw (g), Mt is the total dry mass of straw the period t during decomposition(g), t is the decomposition duration (d).

The straw was crushed and put through a 0.15 mm screen for spectroscopic analysis. Fourier transform infrared spectroscopy (FTIR), scanning electron microscopy (SEM), and elemental analyzer (EA) were used to determine the functional group, morphological structure, and the contents of C, H, O, N and S elements in the straws, respectively.

Greenhouse gas sampling

The CH₄ concentrations were measured by a GC [Agilent Technologies 7890A, Agilent CP7518: 200 °C (50 cm × 530 μm × 10 μm)] coupled with an FID detector. The quantity of injections was 200 μL, with the measurement period of 8 min. The concentration of N₂O was analyzed by GC (Shimadzu GC-2010 plus, GS-CARBONPLOT (30 m × 0.32 μm × 3.00 μm)) equipped with an ECD detector. The quantity of injections was 200 μL, with the measurement period of 10 min.

Soil sample

A portable pH meter was used to determine the soil pH at a 1:2.5 soil-to-water ratio. The concentration of soil organic matter (SOM) was determined utilizing the K₂Cr₂O₇–H₂SO₄ oxidation approach. Dissolved organic carbon (DOC) was extracted using deionized water and analyzed with a TOC analyzer (Shimadzu TOC–VWP, Japan). An AA3 flow analyzer (BRAN + LUEBBE, Germany) was used to measure soil contents of NH₄⁺–N and NO₃⁻–N (Gong et al,. 2023).

Soil enzyme activities and microbial diversity

On days 5, 30, and 90 of the incubation experiment, soil samples were taken and kept at − 20 °C for the study of enzyme activity. Cellulase activity was measured by anthrone colorimetric method [22]. β-Glucosidase activity was measured by nitrophenol colorimetry [22, 23]. Xylanase activity was measured by 3, 5-dinitrosalicylic acid colorimetric method (DNS) [24]. Laccase activity was measured by 2, 2′-azide-bis (3-ethylbenzothiazole-6-sulfonic acid) (ABTS) as substrate [25]. The enzyme activities were measured with a multifunctional microplate reader (Tecan Spark 20 M, Austria).

For microbial communities analysis, soil samples were collected on days 5 and 90 of the culture experiment and kept at − 80 °C. DNA was isolated from 0.5 g soil samples employing a DNA extraction kit (E. Z. N. A Soil DNA Kit, Omega, USA) following the manufacturer’s procedure. The purity of the purified DNA was determined using a Puc-T TA cloning kit (CWBIO, Beijing, China), and the purified DNA was electrophoresed on a 1% (w/v) agarose gel. Primers for 16S rRNA gene amplification were used. The primers for fungal sequencing were ITS5 (5′-GGAAGTAAAAGTCGTAACAAGG-3′) and ITS2 (5′-GCTGCGTTCTTCATCGATGC-3′). Bacterial sequencing primer sequences were 338F (5′-ACTCCTACGGGAGGCAGCA-3′) and 806R (5′-GGACTACHVGGGTWTCTAAT-3′). Following amplification, agarose gel electrophoresis (1.0%) was utilized to confirm the amplified products. The polymerase chain reaction (PCR) protocols were carried out using Phusion^® High-Fidelity PCR Master Mix (New England Biolabs). The DNA library was then constructed and run on the MiSeq Illumina platform at Personalbio Biotechnology Co., Ltd (Shanghai, China).

Data analysis

Statistical analyses were performed using IBM SPSS 22.0, Origin 8.0, and R 4.0.5. Two-way ANOVA was applied to evaluate the effects of pyrolysis temperature and biochar application rate, while one-way ANOVA followed by Duncan’s multiple comparison test (P < 0.05) was used to compare treatments over time. Correlation analyses included Pearson’s and Spearman’s methods, depending on data distribution, with significance levels set at *p < 0.05, **p < 0.01, and ***p < 0.001. Mantel-test heat maps were generated to visualize relationships among variables. Structural equation model analysis was employed to assess the pathways linking biochar, soil chemistry, enzyme activities, microbial populations, straw decomposition, and GHGs. GHGs were recorded from day 0, which was taken as the baseline for subsequent comparisons.

Results

Soil physicochemical properties

The incorporation of rice and maize straw with biochar significantly altered soil chemical properties, including pH, SOM, DOC, NH₄⁺–N, and NO₃⁻–N (Tables S4, S5). Biochar addition increased soil pH relative to the unamended control (5.28 and 6.98), with a rapid rise during the first 30 days followed by stabilization. The extent of alkalinization varied with application rate and pyrolysis temperature. In rice straw treatments, 2.5 and 5% biochar generally produced higher pH values than 10% biochar, whereas in maize straw treatments, pH tended to increase with application rate. SOM contents increased in all treatments, stabilizing after 30 days with rice straw but continuing to accumulate with maize straw. The highest SOM values were observed under RB300-10%, except with 300 °C biochar. DOC exhibited a unimodal pattern, peaking at days 30–60 before declining. In rice straw treatments, DOC was usually highest under 10% biochar, while in maize straw treatments the order at day 30 was MB700 > CK > MB300 > MB500.

NH₄⁺–N declined progressively throughout decomposition, most rapidly in RB300. Conversely, NO₃⁻–N generally increased, although some treatments (e.g., RB500-5%, RB700-2.5%) deviated from this trend. Compared with CK, RB amendments typically maintained higher NH₄⁺–N and lower NO₃⁻–N at day 5, suggesting inhibition of nitrification. In maize straw treatments, NH₄⁺–N patterns were less consistent and showed no clear trend.

Straw decomposition and its components

Straw weight rate loss

Figure 1 displays the straw weight rate loss under various treatments. For the all treatments, the weight loss rate shows a steady increase trend with time, and that of the MB treatments (30.07–33.58%) were higher than the RB treatments (19.64–23.80%). Compared to CK, the straw weight loss rate increased by 12.26–36.04% in RB treatments and 2.92–14.95% in MB treatments, with the highest weight loss observed in RB300-5% and MB500-10% treatments. The degradation rate of straw combined with low pyrolysis temperature biochar was quicker than that with high pyrolysis temperature biochar under the same doses.

Fig. 1

Variation of rice and maize straw weight rate loss (%) with decomposition time

Changes of straw chemical components

The primary components of straw include water-soluble organic substances like sugars, along with insoluble organic substances, including hemicellulose, cellulose, and lignin [26,27,28,29]. In rice straw, hemicellulose and cellulose contents increased initially and then declined, while lignin proportion rose steadily throughout decomposition (Fig. 2a–c). Maize straw showed similar overall trends (Fig. 2d–f), but without a clear decline stage for hemicellulose and cellulose. Compared with CK, both RB and MB treatments significantly accelerated the degradation of hemicellulose, cellulose, and lignin. The strongest effects were observed under low-temperature biochar, particularly RB300 and MB300 treatments, which enhanced fiber decomposition rates by more than 30% in some cases. In contrast, high-temperature biochar (700 °C) slowed the relative decomposition of hemicellulose and cellulose, especially at high application rates (Fig. 2).

Fig. 2

Variation of the specific gravity of hemicellulose (a), cellulose (b) and lignin (c) in rice straw (%); changes in the specific gravity of hemicellulose (d), cellulose (e) and lignin (f) in maize straw (%) with decomposition time

Fig. 3

SEM of rice (a) and maize (b) straw with decomposition time

SEM and FTIR

Based on straw weight loss and chemical composition dynamics, the fastest decomposing treatments were selected (RB300-2.5%, RB300-5%, MB300-10%, and MB500-5%) for SEM and FTIR characterization. SEM images (Fig. 3) showed that biochar addition markedly accelerated structural disintegration. In the early stage, straw exhibited an intact honeycomb-like porous structure. By day 45, biochar-treated straw displayed collapsed pores and loosened fiber bundles, while untreated straw largely maintained its vascular structures. After 90 days, biochar treatments resulted in complete breakdown of the tissue matrix, with extensive voids and a network-like structure, whereas control samples still retained portions of the honeycomb and silicified layers.

Infrared spectrum (Fig. 4) showed that the changes of absorption peaks in rice and maize straws were basically similar during the decomposition process. The strength of the absorption peak at 3430 cm⁻¹ dropped as the straw decomposed, suggesting a reduction in the quantity of amide and carbohydrate molecules. The weakening of the peak of absorption intensity of 2850–2930 cm⁻¹ indicated that the methyl and methylene content of crop straw decreased, that is, the aliphatic compounds content decreased during the decomposition process [30]. An entirely novel absorption peak appeared at 1720–1730 cm⁻¹ (C = O), which was due to the production of unbound ester compounds by crop straw oxidation [31]. The peak at 1600–1610 cm⁻¹ is C = O stretching vibration and amide compounds in lignin linked to aromatic rings. The strength of the absorption peaks at 1150–1160 and 1000–1100 cm⁻¹ decreased in all treatments, suggesting the decomposition of cellulose and hemicellulose [32]. The occurrence of peak at 780–790 cm⁻¹ indicated the transformation of organic silicon to the form of Si–O straw [31]. In addition, Pimentel et al. [33] findings showed that as crop straws decompose, the absorption peaks in infrared spectrum gradually became flat, and some absorption peaks disappear, but the proportion of absorption peaks of difficult compounds increases. This was because the carbohydrates in the plants were easy to be decomposed by micro-organisms, but the substances with aromatic ring structure were difficult to be decomposed. After a long time, aromatic polymer compounds remained and form humic acid.

Fig. 4

Fourier transform infrared spectroscopy of RB (a) and MB (b) treatments with decomposition time

In general, straw with biochar had a higher stretching vibration infrared spectrum than CK. During straw decomposition, the absorption peaks at 1051 or 1032 cm⁻¹ of RB/MB treatments gradually decreased in comparison with CK, but the absorption peaks at 1604 or 1610 cm⁻¹ gradually rose. The peaks intensity ratios of 2920/1640 cm⁻¹ and 1050/1640 cm⁻¹ (Table S3) represent the straw aromatic and aliphatic properties, respectively, which can reflect the degree of straw decomposition [34]. The ratios in the MB and RB treatments were found to be greater than those in the equivalent CK.

Greenhouse gas emissions

Soil CH₄ emissions increased during the early stage of straw decomposition due to the release of labile carbon substrates, then declined as substrate availability decreased (Fig. 5a, b). Compared with CK, both RB and MB treatments initially showed elevated CH₄ fluxes, but emissions were markedly reduced at later stages. For rice straw, RB treatments produced significantly lower CH₄ than CK after day 60, with reductions up to 61.8% by day 90, particularly under RB300-5%. In maize straw treatments, all MB amendments suppressed CH₄ by day 90, with the largest reduction (up to 90.4%) under MB300-10%. These results indicate that low-temperature biochar (300 °C) was most effective in mitigating CH₄ emissions.

Fig. 5

Variation of CH₄ (a, b) and N₂O (c, d) emission fluxes from rice and maize with decomposition time, and the Matel-text correlation analysis between gas emission (CH₄ and N₂O) and environmental factors (addition of straw, biochar, physicochemical properties, and enzyme activities) with rice (e) and maize (f) straw decomposition (*p < 0.05, **p < 0.01, ***p < 0.001)

N₂O emissions increased over time in both rice and maize soils (Fig. 5c, d). However, biochar addition consistently reduced N₂O relative to CK, by 33.9–66.4% in RB treatments and 1.0–91.1% in MB treatments, demonstrating a significant mitigation effect.

Soil enzymes activities

Fiber-degrading flora produces cellulase, which plays a significant role in the straw decomposition. Its activity reflects the ability to convert carboxymethyl cellulose (CMC) into glucose and other small molecules of reducing sugars within a unit of time. As demonstrated in Fig. 6a, b, all the treatments showed an overall upward trend from 0 to 90 days. As a result, cellulase activity correlates favorably with SOM contents. Most RB treatments with 2.5% and 5% addition demonstrated stronger cellulase activity than the CK. Meanwhile, the addition of RB300-10% reduced the cellulase activity in soil with the exception of RB300-2.5%. While maize straw decomposes at a much faster pace than rice straw, the aforementioned phenomena did not occur in the MB treatments, with the exception that adding 5% biochar boosted cellulase activity. The cellulase activity of MB treatments with different pyrolysis temperatures also lacked consistency.

Fig. 6

Variation of the activities of cellulose (a, b), β-glucosidase (c, d), xylanase (e, f), laccase (g, h) from rice and maize straw with decomposition time (p < 0.05)

Xu et al. [35] identified β-glucosidase, xylanase, and laccase as the enzymes responsible for degradation of cellulose, hemicellulose, and lignin. For the rice straw treatments (Fig. 6c), the soil β-glucosidase activity gradually increased with time in all treatments. Compared with CK, the RB treatments had higher soil β-glucosidase activity, expect for 2.5–500 and 10–500 in 5 days. For the maize straw treatments (Fig. 6d), soil β-glucosidase activity first increased and subsequently dropped throughout decomposition, peaking at 30 days except for 2.5–700 and 10–700. Moreover, the soil β-glucosidase activities in MB700 treatments were generally lower than those in MB300 and MB500 treatments in 30 and 90 days. The activity of acid xylanase enzyme showed an upward trend during the rice straw decomposing (Fig. 6e). In comparison with CK, there was no major variation in acid xylanase enzyme activity between the RB treatments after 5 days. Higher enzyme activity in RB treatments was observed from the 30th day and became generally significant across treatments at 90 days. Except for 5–500 in 30 days, the acid xylanase enzyme activity in the MB treatments was noticeably greater than in the CK treatment, as seen in Fig. 6f. In addition, the laccase activity in the RB treatments was lower than the CK except for 10–500, 2.5–700 and 10–700 in 90 days; however, the MB treatments had much greater laccase activity than the CK treatments (Fig. 6g, h).

Microbiological community

At the phylum level, Proteobacteria and Actinobacteriota dominated bacterial communities in rice straw soils, while Proteobacteria were most abundant in maize straw soils (Fig. 7a). Fungal communities were overwhelmingly dominated by Ascomycota in both straw types (> 75%, Fig. 7b).

Fig. 7

Variation of the abundances of soil bacteria (a) and fungi (b) and Alpha diversity (c, d) with decomposition time

Biochar addition significantly influenced microbial richness and diversity (Fig. 7c, d). In rice straw soils, RB300-2.5% treatments increased bacterial diversity and enhanced the relative abundance of Proteobacteria, Actinobacteriota, and Firmicutes, while reducing Acidobacteriota. Conversely, in maize straw soils, biochar lowered Proteobacteria abundance but increased Acidobacteriota, shifting communities from r-strategists to k-strategists.

Fungal diversity also responded differently between the two soils. In rice straw soils, RB300-2.5% treatments increased Shannon and Simpson indices and further enriched Ascomycota, whereas in maize straw soils, MB300 and MB500 treatments reduced fungal diversity and lowered the dominance of Ascomycota.

Overall, biochar reshaped microbial community composition, promoting copiotrophic groups in rice straw systems and favoring oligotrophic groups in maize straw soils, thereby altering decomposition dynamics.

Correlation between straw decomposition/GHGs and environmental factors

Structural equation modeling was used to identify the pathways linking biochar, soil properties, enzymes, and microbial activity to straw decomposition (Fig. 8). For rice straw, decomposition was mainly promoted by biochar and enzyme activities, with biochar strongly enhancing β-glucosidase, xylanase, and cellulase, thereby facilitating hemicellulose and cellulose degradation. In contrast, fungal and bacterial abundances were negatively associated with biochar. For maize straw, decomposition was strongly driven by soil chemical properties and bacterial activity, with biochar indirectly promoting decomposition by improving soil pH and SOM. Fungal activity showed a weak negative correlation with decomposition. Structural equation modeling analysis indicates that biochar primarily accelerates rice straw decomposition through stimulating enzyme activities, whereas maize straw decomposition is more influenced by soil properties and bacterial dynamics.

Fig. 8

Structural equation model analysis of biochar, fungi, bacterial, soil enzyme activities, soil physicochemical properties on straw decomposition. The solid arrows represent significant correlation, where red represents a significant positive correlation and blue represents a significant negative correlation, respectively (*p < 0.05, **p < 0.01, ***p < 0.001)

Mantel-test analysis revealed clear linkages between GHGs and soil variables (Fig. 5e, f). During rice straw decomposition, CH₄ emissions were positively associated with straw addition and β-glucosidase activity, whereas N₂O emissions were negatively correlated with NO₃⁻–N and NH₄⁺–N, but positively with DOC. Biochar application showed strong positive associations with soil pH, SOM, cellulase, and β-glucosidase. In maize straw decomposition, CH₄ emissions were primarily linked to β-glucosidase, while biochar correlated positively with SOM and xylanase, and negatively with NH₄⁺–N. Overall, these results suggest that the influence of biochar on straw decomposition and GHG mitigation was mainly mediated through soil pH, SOM, and the activities of cellulase, β-glucosidase, and xylanase.

Overall, the optimal biochar application for promoting straw decomposition while mitigating greenhouse gas emissions was determined to be a 2.5–5% addition rate combined with a pyrolysis temperature of 300 °C, as evidenced by the highest straw weight loss, fastest fiber component degradation, and most significant reductions in CH₄ and N₂O emissions across both rice and maize straw treatments.

Discussion

Biochar enhances the decomposition rate of straw

Straw decomposition typically occurs in two stages: a rapid phase dominated by soluble sugars, proteins, and hemicellulose degradation, followed by a slower phase constrained by recalcitrant components such as lignin (Fig. 4). In this study, biochar addition significantly accelerated decomposition across both stages, with the most pronounced effect observed under low-temperature biochar (300 °C), which promoted the breakdown of cellulose, hemicellulose, and lignin as well as overall weight loss (Figs. 1 and 2). These findings confirm that biochar not only accelerates carbon turnover but also stimulates the formation of organic macromolecules, consistent with previous reports [36].

Mechanistically, biochar modified soil conditions, elevating pH, increasing DOC, and influencing nitrogen transformation (Tables S4–S6; Sect. "Soil enzymes activities"). The increase in pH enhanced microbial activity and accelerated the release of soluble organics and nutrients, which were incorporated into soil organic matter and sustained microbial metabolism [36, 37]. Adjustments to the C/N balance through nitrogen supply (Tables S4, S5) further promoted fiber degradation [15, 16, 38]. Biochar may also inhibit ammonia-oxidizing or nitrifying bacteria, thereby delaying NH₄⁺–N conversion, which could be explained by inorganic N sorption on biochar surfaces and by its influence on microbial N cycling processes [39, 40].

Our results also showed that enzyme activities (cellulase, β-glucosidase, and xylanase) were strongly correlated with biochar application rates, but less affected by pyrolysis temperature (Tables S6–S9; Fig. 6). This supports the idea that biochar enhances microbial metabolism and enzymatic processes [41]. Biochar demonstrated significant capacity to modulate soil microenvironments and restructure microbial communities, thereby enhancing microbial metabolic activities and enzymatic processes. This might occur due to the following factors: (i) biochar’s porous structure provides habitats and nutrient adsorption sites that promote enzyme efficiency [42]; (ii) surface functional groups interact with soluble organics from straw, reducing enzyme inhibition [43]; and (iii) adsorption of oligomers on biochar stabilizes enzyme binding sites [44]. However, excessive biochar may have inhibitory effects due to enzyme oversaturation, conformational changes, or altered aeration and moisture regimes [45,46,47].

Inhibitory effects were particularly evident at higher pyrolysis temperatures (> 500 °C). Potential causes include the presence of heavy metals and PAHs [48]. As well as enhanced substrate and enzyme adsorption by more stable biochars [49,50,51]. Biochar can also act as an electron acceptor, promoting microbial organic matter degradation [52, 53]. Persistent free radicals may mediate electron transfer and stimulate iron–carbon co-metabolism, thereby facilitating lignin degradation [54]. Under anaerobic conditions, biochar may further promote direct interspecies electron transfer (DIET), enhancing conversion of cellulose metabolites by methanogens and indirectly accelerating decomposition [54].

Overall, these findings demonstrate that biochar significantly enhances straw decomposition by altering soil pH, C/N balance, and enzyme activities, while simultaneously reshaping microbial communities. The positive effects are strongest under low-temperature biochar, whereas high-temperature biochars may introduce inhibitory factors.

Biochar influences the structure of soil microbial communities

Biochar provides a stable matrix that supports microbial diversity by offering protective habitats (Fig. 7). Our results showed that biochar significantly altered microbial community composition during straw decomposition (Fig. 8), but the magnitude of its effect differed between rice and maize straw. Specifically, biochar had a stronger promoting effect on microbial shifts in rice straw treatments, consistent with the higher decomposition rates observed. The addition of biochar favored copiotrophic (r-strategist) groups, such as Proteobacteria, Actinobacteriota, and Firmicutes, which thrive in nutrient-rich environments and produce hydrolytic enzymes that accelerate litter decomposition [41, 51, 55,56,57]. Ascomycota also increased, consistent with their ability to depolymerize fresh organic matter [58]. In contrast, oligotrophic (k-strategist) groups such as Acidobacteriota declined as nutrient availability increased, reflecting competition with rapidly growing r-strategists [55, 59,60,61,62].

Biochar addition to rice straw treatments enhanced soil nutrient content (Table S4), thereby stimulating r-strategists and accelerating microbial metabolism [41]. In maize straw treatments, however, the higher C/N ratio and more complex composition (Table S2) limited labile carbon availability. Under these conditions, biochar promoted the dominance of k-strategists, such as Acidobacteriota, which are positively correlated with high soil C/N ratios [63], and which exhibited greater enzyme sensitivity to substrates.

Together, these results indicate that biochar not only restructures soil microbial communities but also shifts the balance between r- and k-strategists in ways that differ between rice and maize straw decomposition systems. This explains why biochar was more effective in stimulating decomposition in rice straw compared to maize straw.

Biochar influences GHGs during straw decomposition

In this study, CH₄ emissions increased during the initial phase of straw decomposition but declined markedly after 30 days with biochar addition (Fig. 5e, f). The short-term increase can be explained by enhanced methanogen activity due to elevated soil pH [64], thereby promoting methane production [65]. Secondly, the use of biochar might enhance CH₄ emissions in the initial phases by causing a rise in organic matter content during straw decomposition, providing a richer substrate for CH₄ synthesis [18]. Bachoon et al. [66] also found that the addition of cellulose into the soil boosted CH₄ emissions. We speculated that the decomposition of cellulose by β-glucosidase provides abundant substrate for methanogens [35], and thus increases CH₄ emissions. The incorporation of crop straw was the input of exogenous carbon, which not only promoted the production of methane but also stimulates the production of methane from soil organic matter [67].

As decomposition progressed, biochar significantly reduced CH₄ emissions compared with straw alone. This reduction was associated with substrate depletion (Tables S4, S5), adsorption of key methanogenic precursors (acetate, H₂/CO₂) by biochar’s porous structure [43, 68, 69], and competition between methanogens and other microbial groups, such as acetogens and sulfate-reducing bacteria. Biochar also facilitated electron transfer processes, favoring Fe(III)-reducing bacteria over methanogens in anoxic environments [52]. In addition, improved porosity and pH shifts may have suppressed methanogen activity while enhancing methanotrophs, further decreasing CH₄ emissions [70].

Straw addition alone increased N₂O emissions, whereas biochar co-application significantly reduced them. This mitigation is likely due to: (i) improved soil aeration restricting denitrifier activity [71, 72]; (ii) biochar nanopores immobilizing available N and reducing its use in nitrification and denitrification [73, 74], and (iii) enhanced electron transfer promoting N₂O reductase activity, increasing conversion of N₂O to N₂ [11, 73]. Importantly, the inhibition of N₂O emissions was stronger with low-temperature biochar and weakened with increasing pyrolysis temperature.

Overall, our results demonstrate that applying 300 °C biochar with straw return not only promotes straw decomposition but also effectively reduces CH₄ and N₂O emissions, highlighting its potential as a sustainable amendment for field application.

Conclusions

This study demonstrated that biochar significantly alters straw decomposition dynamics and associated GHG emissions. Biochar produced using low-temperature (300 °C), at 2.5–5.0% application rates, accelerated rice and maize straw decomposition by modulating microbial community structure and concomitant increase in soil cellulase and β-glucosidase activities. In contrast, biochar produced using high temperature (500–700 °C) showed weaker effects on decomposition. Biochar addition consistently increased soil pH and SOM, while reducing CH4 and N2O emissions up to 90%, compared with straw alone. Structural equation modeling indicated that enzyme activities were the primary drivers of rice straw decomposition, whereas soil chemical properties and bacterial communities played a larger role in maize straw decomposition. These results highlight the distinct mechanisms by which biochar regulate straw decomposition and GHG mitigation across crop types. Overall, our findings suggest that applying biochar produced using low-temperature at moderate rates provides an effective strategy to enhance straw recycling, improve soil quality, and reduce agricultural greenhouse gas emissions.