Biochar Effectively Reduced N₂O Emissions During Heap Composting and NH₃ Emissions During Aerobic Composting

Abstract

The composting process generates considerable amounts of greenhouse gases, presenting challenges for environmental protection. The utilization of biochar not only improves the composting efficiency but also reduces nitrogen (N) loss during composting. This study aimed to examine the impacts of adding biochar on the composting process, gaseous N emissions, and the bacterial community, as well as to clarify the difference between anoxic and aerobic composting. The experiment was conducted with cow dung and corn straw, with four treatments over 45 days: heap composting (C1), heap composting with 10% biochar (BC1), aerobic composting (C2), and aerobic composting with 10% biochar (BC2). The findings showed that adding biochar significantly reduced N₂O emissions during the heap composting, achieving a cumulative emission reduction of 49.51% compared with composting without biochar. Meanwhile, aerobic composting led to a greater decrease in NH₃ emissions, with a cumulative emission reduction of 56.56%. Additionally, there was a marked increase in the abundance of Bacteroidota and Chryseoline. Biochar reduced nitrogen losses, especially N₂O emissions during heap composting and NH₃ emissions during aerobic composting.

1. Introduction

The increasing demand for meat products has significantly accelerated the expansion of the livestock industry. In China, the annual production of animal manure is estimated to be approximately 3.05 billion metric tons [1], and crop straw generation reaches 1.0 billion tons per year [2]. Composting provides a feasible approach for converting livestock manure and crop straw into organic fertilizer, which can be effectively utilized to improve soil organic matter content and nutrient availability [3]. However, the composting process is accompanied by the uncontrolled release of carbon and nitrogen gases, resulting in considerable losses of these essential elements. Nitrogen loss primarily occurs through the emissions of NH₃, N₂O, and N₂, with NH₃ volatilization accounting for 47% to 77% of the total nitrogen loss [4].

Recent studies have demonstrated that the optimization of process parameters [5], the application of additives [6], and the inoculation of microorganisms [7] can effectively mitigate the emission of carbon- and nitrogen-containing gases, thereby enhancing carbon and nitrogen retention during composting. Among these strategies, the use of additives, particularly biochar, has been shown to be highly effective in increasing carbon and nitrogen retention. For instance, Zhou et al. [8] reported that cornstalk biochar could regulate nitrification processes, thereby reducing NH₃ emissions during layer manure composting. Similarly, Wang et al. [9] reported that incorporating biochar into pig manure compost significantly increased total nitrogen retention while reducing NH₃ and N₂O emissions by 46.85% and 87.94%, respectively. Furthermore, Awasthi Mukesh Kumar et al. [10] demonstrated that adding 10% bamboo biochar to sheep manure compost accelerated the mineralization process, shortened the putrefaction period, and minimized carbon and nitrogen losses. Collectively, these findings provide compelling evidence that biochar application effectively reduces NH₃ and N₂O emissions during the aerobic composting process.

Heap composting and aerobic composting are the two most prevalent composting practices that are adopted in livestock farms and within organic fertilizer processing facilities in China [11]. Heap composting is traditionally conducted under naturally ventilated conditions, enabling the production of compost without mechanical disturbance of the pile. This method is widely adopted because of its operational simplicity and minimal equipment requirements. However, the process is characterized by a relatively slow decomposition rate and often necessitates an extended duration (exceeding one month) to achieve a significant reduction in phytotoxicity [5]. Although the application of biochar has been demonstrated to substantially reduce greenhouse gas emissions during aerobic composting processes, its effects on heap composting remain insufficiently understood and warrant further investigation.

Building on previous research that demonstrates the efficacy of biochar in reducing nitrogen loss during aerobic composting, we hypothesized that biochar can similarly mitigate nitrogen loss under anaerobic composting conditions. We aimed to compare the mechanistic differences between these two composting systems using controlled experiments. Therefore, this study sought to elucidate the effects of biochar addition on reducing NH₃ and N₂O emissions during heap composting as well as to explore the underlying mechanisms involved. Furthermore, the process was compared with aerobic composting to identify both its advantages and differences. This study is anticipated to provide practical guidance for nitrogen retention in waste-to-fertilizer processes by incorporating biochar into anaerobic composting. This approach ensures the retention of a high nutrient content while minimizing nitrogen losses, particularly nitrogen losses due to NH₃ volatilization and N₂O emissions.

2. Materials and Methods

2.1. Test Material

Fresh cow manure was collected from a dairy cattle farm located in Wuxue County, Hubei Province. Biochar was acquired from Bolai Eco-Agriculture Technology Co., Ltd. (Beijing, China)., and corn straw was obtained from an adjacent grain-processing facility. The biochar pyrolyzed at 550 °C exhibited a BET specific surface area of 11.36 m²/g and an average pore diameter of 7.01 nm. The composting site was located in an experimental shed behind the Institute of Plant Protection and Soil Fertilizer of the Hubei Academy of Agricultural Sciences in Wuhan, Hubei Province, China, equipped with a rain-sheltered roof and well-ventilated systems. The basic physicochemical properties of the raw materials and additives that were utilized in the composting process are presented in

Table 1. Basic physical and chemical properties of cow manure, corn straw, and biochar in composting.

Feedstock	Total Carbon %	Total N %	C/N	pH	Moisture Content %
Cattle manure	33.12	2.17	15.20	8.77	79.60
Corn Straw	40.35	1.32	30.51	6.44	12.53
Biochar	44.52	0.78	57.08	9.42	2.5

.

2.2. Experimental Design

Four experimental treatments were established: (1) C1, heap composting without additives; (2) BC1, heap composting with 10% biochar; (3) C2, aerobic composting without additives; and (4) BC2, aerobic composting with 10% biochar. Each composting unit consisted of a polystyrene foam container with a total volume of 200 L (dimensions: 75 cm × 56 cm × 47 cm). The initial substrate was prepared by combining fresh cow manure and straw in a 3:1 mass ratio, and the moisture content was adjusted to 65%. The mixture was divided into two equal portions: one portion served as the control (no biochar) and was allocated to six containers (three replicates each for C1 and C2), while the second portion was homogenized with 10% biochar (by total mass) and distributed equally between the remaining six containers (three replicates each for BC1 and BC2). After mixing the initial materials, no turning or forced aeration was performed in the C1 and BC1 treatments, but manual turning was performed on days 7, 15, 24, and 33 in the C2 and BC2 treatments. The experiment lasted for 45 days.

2.3. Sample Collection and Determination

The pile and ambient temperatures were monitored daily at 09:00 and 15:00, with daily averages calculated for analysis. The pile temperature was determined by taking three measurements with a thermometer in the upper, middle, and lower sections of the pile and then averaging the results to establish the overall pile temperature. Compost samples were collected on days 0, 7, 15, 24, and 30 post-homogenization. Samples were collected from three vertical layers of the pile: the top, middle, and bottom. In each layer, three subsamples were taken horizontally at depths of 15, 25, and 45 cm from the side surface of the composting unit. Each sample (approximately 200 g) was subdivided into two aliquots: one aliquot was immediately subjected to analyses of its physicochemical properties, while the other was stored at −80 °C for subsequent analysis with 16S rRNA high-throughput technology. The total carbon (TC) and total nitrogen (TN) contents of oven-dried samples (105 °C to a constant weight) were quantified with an elemental analyzer (Elementar, Frankfurt, Germany) [12]. The pH was measured with a calibrated pH meter after suspension in deionized water in a solid-to-liquid ratio of 1:10 (w/v) [13]. The moisture content was determined gravimetrically by drying fresh samples at 105 °C until mass stabilization.

The gas samples were collected on days 1, 2, 3, 4, 5, 6, 7, 8, 9, 10, 11, 15, 16, 17, 18, 23, and 30 between 9:00 a.m. and 11:00 a.m. NH₃ levels were determined using the boric acid absorption standard acid titration method (for detailed methods, please refer to Mao et al.’s study [14]). N₂O emissions were measured via the static chamber technique, and the concentrations were analyzed using a 7890A gas chromatograph (Agilent, Santa Clara, CA, USA) [15].

2.4. Data Processing and Analysis

Figure plotting was performed using Origin 2021. Statistical analyses were conducted using the IBM SPSS 26.0 software, with significant differences between treatments determined via a T-test and a one-way analysis of variance (ANOVA) (p < 0.05). The 16S rRNA high-throughput technology was applied to scrutinize the variation in bacterial communities with Majorbio (Shanghai, China). The species diversity index was calculated using the Shannon index and the Chao1 index, which represent microbial community richness and diversity, respectively. Redundancy analysis (RDA) was used to analyze the relationships between parameters, including temperature, moisture, pH, TN, TC, NH₃, N₂O, and microbial community structure.

3. Results and Discussion

3.1. Changes in Physical and Chemical Properties During Composting

3.1.1. Temperature, Moisture, and pH

The temperature dynamics in composting systems are driven primarily by microbial metabolic activity and exothermic organic matter decomposition. As shown in Figure 1a, all treatment groups exceeded 55 °C within the initial 10 days, reaching peak temperatures of 69.0–76.5 °C due to microbial respiration. Notably, compared with the controls, BC1 and BC2 (biochar-amended treatments) maintained elevated temperatures (>55 °C) for an extended duration of 13 days. These findings align with those of Wang et al. [16], who reported increased thermal retention in poultry manure compost containing biochar [17]. The prolonged thermophilic phase observed in biochar-amended groups may be attributed to the role of biochar in facilitating thermophilic microbial colonization, which enhances aeration within the pore structure, reduces heat dissipation, and amplifies metabolic heat generation [15]. During days 16–34, the mean temperatures in the anoxic treatment groups (C1: 38.6 °C; BC1: 38.4 °C) remained lower than those in the aerobic treatment groups (C2: 42.4 °C; BC2: 43.3 °C). This disparity was maintained into the cooling phase, where aerobic systems retained more residual heat. We hypothesize that incomplete organic matter degradation during the early thermophilic phase of aerobic composting provided sustained substrate availability for microbial activity, delaying the decrease in temperature. Conversely, anoxic conditions likely accelerated initial organic matter utilization, leading to faster substrate depletion and a more gradual cooling trajectory [18,19].

Moisture content serves as a critical medium for nutrient transport and microbial metabolism in composting systems [20]. As shown in Figure 1b, all the treatments resulted in a progressive decrease in moisture content over time. By the end of the composting period, the moisture reduction reached 23.3% (C1), 23.1% (BC1), 17.9% (C2), and 21.1% (BC2) relative to the initial value. Moreover, the anoxic treatments (C1 and BC1) resulted in significantly greater moisture loss than the aerobic treatments (C2 and BC2). This difference may have arisen due to the prolonged thermophilic phase in anoxic systems (Figure 1a), which intensifies evaporative water loss due to sustained high temperatures. Notably, the biochar-amended treatments (BC1 and BC2) resulted in marginally higher temperature profiles than their respective controls (C1 and C2), further confirming the link between thermal persistence and moisture evaporation. During heap composting, moisture reduction occurred predominantly during the thermophilic phase, whereas aerobic systems experienced gradual water loss during the cooling and maturation phases. The addition of biochar amplified moisture loss in the aerobic treatment groups (BC2: 21.1% vs. C2: 17.9%), likely because the macroporous structure of biochar increases oxygen diffusion and vapor transport [21]. Although the microporous surface of biochar can transiently enhance water retention, its overall porosity may accelerate the dehydration process by facilitating airflow and heat distribution within the matrix. This dual effect highlights the context-dependent influence of biochar on compost moisture dynamics.

All treatments resulted in a triphasic pH trajectory (Figure 1c): an initial increase (pH 8.75–8.87 to 9.19–9.40), followed by a decrease (pH 8.79–9.31), and a subsequent rebound. The pH value increase in the first phase corresponded to ammonia (NH₃) volatilization during nitrogen mineralization. The subsequent pH decrease was driven by reduced NH₃ emissions and the accumulation of low-molecular-weight fatty acids (e.g., acetic and propionic acids) and CO₂ from organic matter decomposition. The final pH increase (9.01–9.32) reflected the microbial mineralization of residual organic acids and the stabilization of alkaline minerals [10,11]. Compared with the control treatments (C1 and C2), the biochar-containing treatments (BC1 and BC2) resulted in increased pH shifts, with final pH increases of +0.46 (BC1) and +0.44 (BC2) units versus +0.27 (C1) and +0.42 (C2). This finding is consistent with the findings of Li et al. [22], who reported a 0.3-unit pH increase in biochar-swine manure–corn stover compost, which was attributed to the inherent alkalinity and buffering capacity of biochar. Similarly, Sánchez-García et al. [23] reported that biochar incorporation increases the pH of compost by adsorbing acidic metabolites and increasing ammonification rates. The pronounced pH modulation in the aerobic treatment groups (BC2 > C2) likely stemmed from the interaction of biochar with oxygen availability, which promoted nitrifier activity and organic acid degradation.

3.1.2. Total Carbon, Total Nitrogen, and C/N Ratio

The temporal variations in the total carbon (TC) content across the four treatment groups are shown in Figure 2a. During the initial 9 days, the TC decreased in all treatment groups, reflecting rapid microbial mineralization of labile organic carbon. However, during days 10–30, the TC increased in the anoxic treatment group (C1: +5.2%; BC1: +6.8%), likely because mass loss from volume reduction and moisture depletion concentrated residual carbon [24]. The TC subsequently steadily decreased in all the systems until composting concluded, with higher carbon loss rates in the thermophilic phase (days 0–15) than in the later stages, which is consistent with the progressive depletion of degradable substrates [25]. The final TC values were 32.7% (C1), 32.3% (C2), 38.6% (BC1), and 38.5% (BC2), representing net reductions of 9.85%, 11.5%, 5.00%, and 4.30% from the initial levels, respectively. The biochar-containing treatments (BC1 and BC2) resulted in significantly lower carbon loss, which is consistent with the findings of Hagemann et al. [26], who attributed the reduced carbon loss in manure compost to the recalcitrant structure and adsorption properties of biochar. The highly specific surface area and cation exchange capacity of biochar facilitate the stabilization of organic carbon via chemical bonding with oxidized functional groups (e.g., carboxyl and hydroxyl groups), thereby increasing carbon sequestration [22]. These results demonstrate the potential of biochar to reduce carbon loss during composting, thereby enhancing both carbon retention and the quality of the final product.

In the control treatment groups (C1 and C2), the TN rapidly increased during the first 17 days, followed by a slow accumulation phase. The biochar-containing treatments (BC1 and BC2) resulted in similar initial increases, peaking on day 17 but stabilizing thereafter (Figure 2b). This pattern is consistent with the concentration effect caused by rapid moisture loss and organic matter mineralization, where mass reduction outpaces nitrogen volatilization, leading to proportional TN enrichment [22]. By the end of composting, the TN had increased by 29.9% in the C1 and C2 groups, whereas it had increased by 17.6% and 17.7% in the BC1 and BC2 groups. Contrary to expectations, biochar incorporation did not improve nitrogen retention, as evidenced by lower TN gains in the BC1/BC2 groups than in the controls. This suggests that the porous structure of biochar may increase ammonia (NH₃) volatilization by increasing oxygen diffusion, offsetting its potential benefits regarding adsorption. Similar findings were reported by Awasthi et al. [10], where biochar alone failed to mitigate nitrogen loss in food waste compost unless it was used alongside microbial inoculants. These results highlight the need to integrate biochar with complementary additives (e.g., zeolites and nitrification inhibitors) to optimize nitrogen conservation in composting systems [27].

The carbon-to-nitrogen (C/N) ratio serves as a critical indicator of compost maturity. Yang et al. [28] reported that a decrease in the C/N ratio from an initial range of 25–35 to 10–15 signifies stabilization and maturation of the compost. In this study, the C/N ratios of all the treatment groups progressively decreased throughout the composting process, reaching final values of 12.3 (C1), 12.2 (C2), 21.9 (BC1), and 22.3 (BC2) after 45 days (Figure 2c). Notably, the C/N ratios of the biochar-amended treatment groups (BC1 and BC2) remained above the threshold of 20, which is conventionally associated with incomplete decomposition [29]. This finding is consistent with a previous report by Liu et al. [17], which suggested biochar addition does not accelerate the degradation of organic matter during composting. A persistently elevated C/N ratio reflects an imbalance in carbon and nitrogen availability, wherein excessive carbon relative to nitrogen limits microbial activity and prolongs the composting period.

3.1.3. Changes in NH₃ and N₂O Emissions During Composting

NH₃ emissions rapidly peaked during days 3–7 following the initiation of composting, followed by a sharp decline to undetectable levels (Figure 3a). Throughout the 45-day process, the heap composting systems (C1 and BC1) presented significantly lower NH₃ emission rates than the aerobic systems (C2 and BC2). The peak emission values reached 6.7 mg m⁻² h⁻¹ (C1), 9.1 mg m⁻² h⁻¹ (BC1), 116.0 mg m⁻² h⁻¹ (C2), and 60.8 mg m⁻² h⁻¹ (BC2). Rapid NH₃ release is strongly correlated with elevated temperatures (Figure 1a) and alkaline pH conditions (Figure 1c), as microbial degradation of organic nitrogen compounds generates free NH₃ under thermophilic and high-pH conditions [30]. Furthermore, periodic pile turning introduces oxygen, exacerbating NH₃ volatilization by destabilizing the NH⁺₄‒NH₃ equilibrium [2]. The cumulative NH₃ emissions totaled 0.88 g m⁻² (C1), 0.63 g m⁻² (BC1), 9.22 g m⁻² (C2), and 4.01 g m⁻² (BC2) (Figure 3b). Relative to the control treatments (C1 and C2), the biochar-amended systems (BC1 and BC2) resulted in 27.8% and 56.6% reductions in NH₃ emissions, respectively. Conversely, aerobic composting (C2 and BC2) increased emissions by 90.4% and 84.2%, respectively, compared with the anoxic systems (C1 and BC1). The mitigation effect of biochar is attributed to its porous structure and high surface area, which increase NH₃ adsorption and immobilization [31]. Critically, the heap composting systems demonstrated substantially lower cumulative NH₃ emissions than the aerobic treatments, which is consistent with studies linking frequent aeration and mechanical disturbances to elevated nitrogen losses [32]. These results highlight that microaerobic conditions in heap composting inherently suppress NH₃ volatilization by limiting oxygen-driven ammonia liberation.

As shown in Figure 3c, N₂O emissions across all the treatment groups followed a biphasic pattern, characterized by an initial rapid increase during the thermophilic phase and a gradual decrease in the later composting stages. The peak emission rates reached 10.7 mg m⁻² h⁻¹ (C1), 12.9 mg m⁻² h⁻¹ (BC1), 31.9 mg m⁻² h⁻¹ (C2), and 21.6 mg m⁻² h⁻¹ (BC2), with the anoxic treatments (C1 and BC1) resulting in a transient peak on day 30 followed by rapid attenuation (Figure 3c). Early-stage N₂O increases are strongly correlated with compost turning events, which disrupt anaerobic microenvironments and stimulate nitrifier–denitrifier activity by enhancing oxygen diffusion, a phenomenon that is consistent with the findings of Huang et al. [2]. Conversely, in anoxic systems, anaerobic microenvironments transiently suppress denitrifying bacterial activity, thereby suppressing N₂O production. During maturation, renewed N₂O emissions following anoxic treatments are attributed to residual denitrification under conditions of decreasing temperatures and elevated nitrite accumulation [33]. The cumulative N₂O emissions totaled 5.99 g m⁻² (C1), 3.02 g m⁻² (BC1), 2.38 g m⁻² (C2), and 2.28 g m⁻² (BC2) (Figure 3d). Biochar amendment reduced the emissions by 49.5% (BC1 vs. C1) and 4.16% (BC2 vs. C2), whereas aerobic systems (C2 and BC2) exhibited emissions that decreased by 60.3% and 24.6% relative to their anoxic counterparts (C1 and BC1). This mitigation is mechanistically linked to the alkaline properties of biochar, which modulate denitrifier enzyme activity [34], and its porous matrix, which suppresses microbial metabolism via thermal insulation during the thermophilic phase. Furthermore, in this anoxic environment, the electron-shuttling capacity of biochar, facilitated by high electrical conductivity and redox-active surface functional groups, promotes microbial N₂O reduction to N₂ [35], thereby decreasing N₂O emissions.

3.2. Microbiological Changes During Composting

3.2.1. Changes in Shannon and Chao1 Indices of Community Diversity

The Chao1 and Shannon indices serve as quantitative measures of microbial community richness and diversity, respectively, within composting systems [36]. Higher Chao1 values denote greater species richness, whereas elevated Shannon indices reflect increased community diversity [37]. Initial assessments revealed that, compared with the control treatments (C1 and C2), the biochar-containing treatments (BC1 and BC2) reduced the Chao1 and Shannon indices, which was likely attributable to the influence of biochar on the initial substrate C/N ratio (Figure 2c). However, at the conclusion of the 45-day composting period, both indices significantly increased in the BC1 and BC2 groups (Figure 4a,b), indicating increased species richness and community diversity in the biochar-amended systems. This shift suggests that biochar initially moderates the microbial community structure but, ultimately, fosters a more diverse and robust microbiota. Mechanistically, biochar enhances microbial nitrogen assimilation capacity by serving as a habitat for microbial colonization and a reservoir for nutrient retention, thereby mitigating nitrogen losses during composting [38]. These findings highlight that both composting methodology (aerobic vs. anoxic) and additive application (biochar) shape bacterial community dynamics and functional outcomes.

3.2.2. Succession of Bacterial Communities at Phylum and Genus Levels

The impact of biochar on the bacterial community composition at the phylum and genus levels is shown in Figure 5. The dominant phyla across all the treatment groups included Bacteroidota, Firmicutes, Proteobacteria, Chloroflexi, Actinobacteria, and Gemmatimonadota, which collectively represented ~90% of the total relative abundance (Figure 5a), and this result is consistent with the microbial community lineages observed in previous studies on composting [6,39]. Firmicutes, which are prevalent during the thermophilic phase, exhibited enhanced abundance in the biochar-amended treatment groups (BC1 and BC2), which was correlated with sustained high temperatures (Figure 1a). This phylum’s resilience under thermal stress is attributed to its spore-forming capacity [40], while its metabolic specialization in carbohydrate and polysaccharide degradation drives organic matter breakdown in aerobic systems [41]. Proteobacteria and Bacteroidota are critical for carbon/nitrogen cycling, with Bacteroidota further facilitating cellulose/hemicellulose decomposition [42,43]. Notably, the abundance of Chloroflexi, which is a metabolically versatile phylum with anaerobic predominance, increased in the biochar treatment groups (BC1 and BC2), likely because the porous structure of biochar increases microaerobic niches and oxygen diffusion [44]. By day 45, Bacteroidota dominated all the treatment groups, constituting 49.4% (C1), 33.1% (BC1), 36.2% (C2), and 38.0% (BC2) of the communities (Figure 5a). This phylum’s late-stage prominence highlights its role in terminal organic mineralization and nutrient cycling [42]. The selective enrichment of Firmicutes and Chloroflexi by biochar highlights the dual function of biochar as a microbial habitat and physicochemical modulator, shaping community succession to optimize decomposition efficiency.

The dominant bacterial genera across the composting treatment groups are detailed in Figure 5b. Unclassified genera collectively constituted 40–60% of the microbial community, which is consistent with prior metagenomic studies of composting systems [45]. Notably, Chryseolinea presented a higher relative abundance in the aerobic treatments (C2 and BC2), whereas Flavobacterium dominated in the anoxic systems (C1 and BC1). These genera are hypothesized to contribute to the observed mitigation of carbon and nitrogen emissions, as Chryseolinea is associated with lignocellulose degradation [46], and Flavobacterium participates in nitrogen immobilization via protease activity [47]. Anseongella—undetectable in pre-composting samples—emerged during maturation, reaching a relative abundance of 3–6% in all the treatment groups by day 45. This genus is associated with aromatic compound metabolism and humic acid synthesis [48], suggesting its role in accelerating humification. The delayed proliferation of Anseongella was consistent with the progression of stabilized organic matter formation, highlighting its potential as a biomarker for compost maturity.

3.3. Correlation of Microbial Community Structure with Environmental Factors

Redundancy analysis (RDA) was performed to elucidate the mechanistic interplay between physicochemical properties, gaseous emissions, and microbial communities during composting (Figure 6). At the phylum level, environmental parameters (e.g., pH, TN, TC, the C/N ratio, and temperature) and gas emissions (NH₃ and N₂O) collectively explained 57.11% of the bacterial community variation (Figure 6a). Bacteroidota was strongly positively correlated with pH, TN, NH₃, and N₂O, which is consistent with its role in nitrogen transformation and organic acid metabolism under alkaline conditions [12]. Conversely, Firmicutes—predominant during the thermophilic phase—showed significant negative correlations with NH₃ and N₂O emissions, suggesting that biochar has the capacity to mitigate emissions by enriching this heat-resistant, carbohydrate-degrading phylum [45]. The relative abundance of Chloroflexi was positively correlated with TC content, C/N ratio, and temperature, likely due to its metabolic versatility in carbon-rich, thermophilic environments, whereas the relative abundance of Proteobacteria displayed inverse trends, potentially reflecting its sensitivity to oxygen fluctuations [42].

At the genus level, axes 1 and 2 accounted for 59.26% of the microbial variation (Figure 6b). The relative abundance of Chryseolinea was tightly linked to elevated pH and NH₃ levels, which is consistent with its reported role in ammonification [46], whereas the abundance of Flavobacterium was correlated with TN and N₂O levels, implicating it in denitrification pathways. These findings validate the functional divergence inferred from the emission profiles (Figure 3d). Importantly, biochar modulates composting dynamics by altering key environmental drivers (e.g., pH and oxygen availability), which subsequently reshape the microbial community structure and metabolic activity. This cascade effect optimizes carbon and nitrogen cycling, reducing gaseous losses while promoting humification.

4. Conclusions

This study demonstrated that biochar amendment enhances composting efficiency by prolonging the thermophilic phase and raising the pH, thereby improving organic matter stabilization and product safety. Specifically, biochar-added heap composting (BC1) reduced NH₃ and N₂O emissions by 27.8% and 49.5%, respectively, compared with conventional heap composting (C1). While aerobic composting with biochar (BC2) reduced cumulative NH₃ emissions by 56.6% relative to the control, BC1 outperformed BC2 by suppressing NH₃ emissions by 84.2%. These reductions were mechanistically attributed to biochar’s microbial modulation: it suppressed Bacteroidota (associated with ammonia volatilization) and enriched Chryseolinea (linked to nitrogen retention), restructuring microbial networks to mitigate gaseous losses. Moreover, adding 10% biochar during initial aerobic composting reduced ammonia volatilization and nitrogen loss while enhancing compost quality. At the farm level, a 45-day co-composting process of livestock manure with straw and 10% biochar not only lowered N₂O emissions but also improved sanitization.