Nitrogen Transformation Mechanisms and Compost Quality Assessment in Sustainable Mesophilic Aerobic Composting of Agricultural Waste

Abstract

This study examines nitrogen transformation mechanisms and compost quality in mesophilic aerobic composting of wheat straw, utilizing cow manure as a co-substrate to promote sustainable agricultural waste management. Two composting systems were established: group A (control) and group B (10% cow manure addition by wet weight). The addition of cow manure accelerated early organic matter decomposition and increased total nitrogen retention in group B. Nitrogen losses occurred primarily via ammonia volatilization during the initial and final composting stages, while functional gene analysis revealed enhanced ammonification and nitrification in both systems. Microbial community analysis showed that cow manure addition promoted nitrogen-fixing bacteria in the early phase and fungi associated with complex organic degradation in later stages. These findings underscore the potential of cow manure to enhance compost maturity, improve nitrogen efficiency, and support the development of sustainable composting practices that contribute to resource conservation.

1. Introduction

As a major agricultural nation, China generates a large volume of waste each year, including crop residues, bagasse, and corncobs, most of which remain underutilized. These organic solid wastes are compositionally complex, rich in organic matter (OM) and essential plant nutrients (N, P, and K), but also harbor pathogenic microorganisms, toxic pollutants, and heavy metals [1,2], which are harmful to environmental sanitation. Therefore, the effective management and utilization of agricultural waste as a resource is a critical global issue that urgently needs to be addressed.

Composting offers an environmentally friendly and cost-effective solution for the decomposition of organic waste [3]. Mature compost, being stable, safe, and nutrient-rich (e.g., humic acid, available nitrogen), can be applied as an organic fertilizer and soil amendment. However, variations in decomposition rates, driven by factors like aeration, temperature, and the carbon-to-nitrogen (C/N) ratio, have spurred the development of diverse methods, including aerobic, anaerobic, mesophilic, and thermophilic composting. Although aerobic composting is widely used for recycling and treating organic solid waste, significant nutrient losses—particularly nitrogen (45–75%)—during high-temperature composting remain a major challenge, diminishing the compost’s fertilizer value and its competitiveness for land application [4]. Extensive studies have shown that ammonia (NH₃) volatilization is the primary cause of nitrogen loss during high-temperature aerobic composting, accounting for 79–94% of total nitrogen (TN) loss [5], while nitrogen lost as nitrous oxide (N₂O) constitutes approximately 0.2–9.9% [6]. During the composting process, most nitrogen loss occurs in the thermophilic phase, estimated to account for 40–70% of the initial nitrogen content [7]. Extensive research on nitrogen transformation and loss during high-temperature aerobic composting has identified dynamic shifts in microbial community structure and function—driven by factors such as elevated temperatures [8], high pH levels [9], and frequent turning [10]—as the primary contributors to nitrogen loss. Building on this understanding, most current studies focus on nitrogen retention strategies. Approaches such as adding exogenous materials (e.g., zeolite, magnetite, amino acids, biochar, amoxicillin) [3,11,12,13,14], inoculating nitrogen-fixing microbial agents (e.g., Bacillus licheniformis, Gordonia paraffinivorans N52, Penicillium chrysogenum) [15,16,17], using composting with a membrane cover [18], or co-composting [19] have been proposed to enhance nitrogen adsorption, nitrogen assimilation, nitrification, and nitrogen fixation pathways, while reducing enzyme activity and gene abundance related to denitrification to mitigate nitrogen loss. However, research on nitrogen transformation and loss mechanisms during mesophilic composting remains largely unexplored. Only a few studies have preliminarily investigated aspects such as OM transformation [20], microbial succession (Proteobacteria → Actinobacteria) [21], and the production of high-quality substrates (with seedling survival and germination indices of 87.0% and 161%, respectively) [22]. It has also been demonstrated that adding 2–3% calcium cyanamide can meet sanitation standards without significantly inhibiting native microbial populations or organic substrate degradation [23]. Moreover, there is an urgent need for in-depth research on nitrogen transformation and loss mechanisms in mesophilic aerobic composting, focusing on the potential trade-off between extended maturation time, energy savings, and nutrient retention. This research is crucial for promoting the adoption of this technology in rural, remote, and scenic areas.

During the aerobic composting of agricultural waste, microbial communities undergo a series of shifts in response to changes in the physicochemical properties of the compost, mediating nitrogen transformation and influencing nitrogen loss. We hypothesize that the addition of cow manure, as a conditioning agent, influences the fungal and bacterial community structures as well as the carbon and nitrogen metabolic pathways and functions within the mesophilic aerobic composting system of wheat straw. This, in turn, may delay nitrogen loss, offering profound implications for the development of nitrogen retention technologies, advancing our understanding of microecological mechanisms in mesophilic aerobic composting. The objectives of this study were as follows: (1) to elucidate the nitrogen transformation processes and quantify nitrogen loss during mesophilic aerobic composting of straw, and to evaluate compost quality; (2) to uncover the succession patterns of microbial communities during mesophilic aerobic composting; and (3) to clarify the potential mechanisms and pathways of nitrogen loss in mesophilic aerobic composting.

2. Materials and Methods

2.1. Composting Design

Based on preliminary experimental experience, wheat straw collected from farmland in Funan County (Anhui Province, China) was used as the primary substrate to simulate agricultural solid waste composting. A lab-scale aerobic composting experiment was conducted using open fermentation bins, with two composting systems (group A and group B), each with three replicates. The fermentation bins were 150 L plastic barrels, with evenly distributed 3–5 mm holes at the bottom to facilitate liquid drainage and oxygen intake during composting. Wheat straw was cut into 1–2 cm segments for use. Fresh soil from a depth of 0–10 cm was sieved through a 10-mesh screen to remove stones, plant debris, and other impurities. Wheat straw and fresh soil were thoroughly mixed in a 1:1 mass ratio, with market-purchased soybean meal added to adjust the carbon-to-nitrogen (C/N) ratio of the composting substrate to an initial value of 25 [12]. The physicochemical characterization of each raw material is listed in Table S1. The mixed substrate was then divided into six equal portions using the quartering method, with each portion weighing 19.5 kg wet. Next, fresh cow manure was added to group B at a rate of 10% of the wet weight of the substrate, while sterile ultrapure water was added to group A at a rate of 6% of the wet weight. In the early stages of composting, the material was turned every 2 days, while in the later stages, turning occurred every 7 days.

2.2. Sample Collection and Processing

Samples were collected at appropriate times based on the actual temperature changes in the compost. A multi-point sampling method was used, with approximately 50 g of material taken from both the center and the periphery of the compost pile. The samples were divided into two portions: one 100 g fresh sample was dried at 25 °C, ground, sieved, and collected in a sealed bag for storage in a dry indoor environment; the other 100 g sample was stored at −80 °C for future analysis.

2.3. Determination of Basic Index of Compost Sample

Temperature measurements were taken three times daily (morning, noon, and evening) using a probe thermometer (YH-101, Yuan Hao, Haofei Instrument Technology Co., Ltd, Dongtai City, China). The moisture content was determined using the drying method. Precisely 5 g of the sample, sieved through an 18-mesh screen, was weighed and mixed with distilled water at a water-to-soil ratio of 10:1 (v/m). The mixture was stirred for 1 min using a glass rod to ensure full dispersion of the sample, then allowed to settle for 0.5 to 1 h. The pH and EC values of the filtrate were measured using a portable water quality meter (Orion Star A329, Thermo Scientific, Waltham, MA, USA). Total OM was determined by calculating the mass difference before and after burning in a muffle furnace at 550 °C for 7 h [12]. Total carbon and total organic carbon were measured using an automated elemental analyzer (APSA-370, Yanaco Co., Ltd., Kyoto, Japan). The measurement of total nitrogen (TN), ammonia nitrogen (NH₃-N), nitrate nitrogen (NO₃⁻-N), and nitrite nitrogen (NO₂⁻-N) were conducted according to Li et al. [24]. The concentrations of five heavy metals (As, Cd, Pb, Cr, and Hg) in the compost samples were determined using inductively coupled plasma mass spectrometry (ICP-MS 7500a, Agilent Technologies, Santa Clara, CA, USA).

The ammonia volatilization rate (F, mg/(m²·h)) was measured using the double-layer sponge aeration method [25]. Two sponges were each soaked in 15 mL of a phosphoric acid–glycerol solution. One sponge was placed at the top of a PVC tube to prevent interference from external gases, while the other was positioned a certain distance from the bottom of the tube to absorb ammonia volatilized from the compost. After placing the PVC tube on the compost pile for 1 h, the lower sponge was removed and shaken in 1 mol/L KCl solution for 1 h to extract the absorbed ammonia. The concentration of NH₃-N in the extract was then measured. The ammonia volatilization rate was calculated using Equation (1).

$$\mathrm{F}={\displaystyle \frac{\mathrm{C}\mathrm{o}\mathrm{n}\mathrm{c}\mathrm{e}\mathrm{n}\mathrm{t}\mathrm{r}\mathrm{a}\mathrm{t}\mathrm{i}\mathrm{o}\mathrm{n} \mathrm{o}\mathrm{f} \mathrm{a}\mathrm{m}\mathrm{m}\mathrm{o}\mathrm{n}\mathrm{i}\mathrm{a} \mathrm{n}\mathrm{i}\mathrm{t}\mathrm{r}\mathrm{o}\mathrm{g}\mathrm{e}\mathrm{n} \mathrm{i}\mathrm{n} \mathrm{t}\mathrm{h}\mathrm{e} \mathrm{e}\mathrm{x}\mathrm{t}\mathrm{r}\mathrm{a}\mathrm{c}\mathrm{t}\left(\mathrm{m}\mathrm{g}/\mathrm{L}\right)\times \mathrm{V}\mathrm{o}\mathrm{l}\mathrm{u}\mathrm{m}\mathrm{e} \mathrm{o}\mathrm{f} \mathrm{t}\mathrm{h}\mathrm{e} \mathrm{e}\mathrm{x}\mathrm{t}\mathrm{r}\mathrm{a}\mathrm{c}\mathrm{t}\left(\mathrm{L}\right)}{\mathrm{C}\mathrm{r}\mathrm{o}\mathrm{s}\mathrm{s}-\mathrm{s}\mathrm{e}\mathrm{c}\mathrm{t}\mathrm{i}\mathrm{o}\mathrm{n}\mathrm{a}\mathrm{l} \mathrm{a}\mathrm{r}\mathrm{e}\mathrm{a}\left({\mathrm{m}}^2\right)\times \mathrm{M}\mathrm{e}\mathrm{a}\mathrm{s}\mathrm{u}\mathrm{r}\mathrm{e}\mathrm{m}\mathrm{e}\mathrm{n}\mathrm{t} \mathrm{t}\mathrm{i}\mathrm{m}\mathrm{e}\left(\mathrm{h}\right)}}$$

(1)

Chinese cabbage seeds and amaranth seeds were used for germination index (GI) testing, which is sensitive to phytotoxic substances and has a fast germination rate. Compost samples were extracted with deionized water at a solid-to-liquid ratio of 1:10 (w/v), shaken for 1 h at room temperature, and the extract was filtered. A sterile Petri dish was lined with filter paper, and 20 Chinese cabbage seeds and amaranth seeds were evenly distributed in each dish. Then, 5 mL of the extract was added, with distilled water used as a blank control. The dishes were incubated in a dark incubator at 20 ± 2 °C for 48 h. Seed germination rate and root length were recorded, and germination rate was calculated using Equation (2) [12].

$$GI={\displaystyle \frac{\mathrm{S}\mathrm{e}\mathrm{e}\mathrm{d} \mathrm{g}\mathrm{e}\mathrm{r}\mathrm{m}\mathrm{i}\mathrm{n}\mathrm{a}\mathrm{t}\mathrm{i}\mathrm{o}\mathrm{n} \mathrm{o}\mathrm{f} \mathrm{t}\mathrm{r}\mathrm{e}\mathrm{a}\mathrm{t}\mathrm{m}\mathrm{e}\mathrm{n}\mathrm{t} \ast \mathrm{R}\mathrm{o}\mathrm{o}\mathrm{t} \mathrm{l}\mathrm{e}\mathrm{n}\mathrm{g}\mathrm{t}\mathrm{h} \mathrm{o}\mathrm{f} \mathrm{t}\mathrm{r}\mathrm{e}\mathrm{a}\mathrm{t}\mathrm{m}\mathrm{e}\mathrm{n}\mathrm{t}\left(\mathrm{m}\mathrm{m}\right)}{\mathrm{S}\mathrm{e}\mathrm{e}\mathrm{d} \mathrm{g}\mathrm{e}\mathrm{r}\mathrm{m}\mathrm{i}\mathrm{n}\mathrm{a}\mathrm{t}\mathrm{i}\mathrm{o}\mathrm{n} \mathrm{o}\mathrm{f} \mathrm{w}\mathrm{a}\mathrm{t}\mathrm{e}\mathrm{r} \ast \mathrm{R}\mathrm{o}\mathrm{o}\mathrm{t} \mathrm{l}\mathrm{e}\mathrm{n}\mathrm{g}\mathrm{t}\mathrm{h} \mathrm{o}\mathrm{f} \mathrm{w}\mathrm{a}\mathrm{t}\mathrm{e}\mathrm{r}\left(\mathrm{m}\mathrm{m}\right)}}\ast 100\mathrm{\%}$$

(2)

2.4. High-Throughput Sequencing

For convenience, these samples were marked as A0, B0, A31, B31, A66, and B66, respectively. The bacterial 16S rDNA was amplified using primers 515F (5′-GTGCCAGCMGCCGCGGTAAT-3′) and 806R (5′-GGACTACHVGGGTWTCTAA-3′), and the methods of measurement followed Yang et al. [26]. The fungal 18S rDNA was amplified using primers ITS1F (5′- CTTGGTCATTTAGAGGAAGTAA-3′) and ITS2R (5′- GCTGCGTTCTTCATCGATGC-3′), and the methods of measurement followed Wen et al. [27]. The functional annotation was predicted by PICRUSt2, utilizing the Kyoto Encyclopedia of Genes and Genomes (KEGG) database.

2.5. Statistical Analysis

Statistical analysis was conducted in Excel 2021 (Microsoft Corp., Redmond, WA, USA) and Origin 2017 64-bit (Origin Pro Lab Corp., Northampton, MA, USA). The pheatmap package in R (version 4.2.2) was utilized to conduct the bacterial community heatmap. Network analysis was crafted using Gephi 0.9.4, adhering to Spearman’s coefficients. Experimental data were visualized using Origin 2017 64-bit software (Origin Pro Lab Corp., Northampton, MA, USA).

3. Results and Discussion

3.1. Physicochemical Properties of Compost

The whole composting process lasted for 66 days. The dynamic changes of temperature in group A and group B are shown in Figure 1a. In the experimental group B, the temperature rose rapidly at the start of composting, reaching a peak of 30.3 °C on day 3, then fluctuated downward before stabilizing around 25 °C. This suggests that the addition of cow manure promoted the growth of microorganisms responsible for early-stage OM degradation [28], thereby accelerating the decomposition process. In contrast, the temperature in control group A rose gradually, reaching a peak of 28.73 °C only in the mid composting phase (day 28). Although temperatures in both systems did not exceed 31 °C throughout composting, they remained above ambient temperature, indicating that straw decomposition in mesophilic aerobic composting occurred gradually over time, with cow manure addition accelerating this process.

Moisture content is crucial for maintaining and regulating microbial activity, temperature, and oxygen supply during aerobic composting. Studies indicate that an initial moisture content of 55–65% optimizes microbial activity and compost maturity [29]. Figure 1b shows that the initial moisture contents of control group A and experimental group B were 64.61% and 65.13%, respectively, aligning with the optimal conditions for aerobic composting. Although water was added during the turning process, moisture content in both systems consistently decreased throughout the composting period, with final moisture levels of only 60.34 ± 4.99% for group A and 51.55 ± 9.07% for group B.

The final pH of compost is widely used to assess the quality of compost products due to its influence on soil pH and the bioavailability of nutrients to plants post-application [13]. pH significantly affects microbial activity during the composting process, with many studies indicating that optimal pH levels for efficient composting range from 8.0 to 9.0 [30]. The pH changes in both piles were basically the same, which increased rapidly to over 8.0 at the initial composting phase and were finally maintained between 8.3 and 8.5 (Figure 1c). This might be due to the decomposition of organic acids and the ammonification of nitrogenous OM [31]. In experimental group B, the pH reached its maximum of 8.71 on day 17, followed by a decline to a minimum of 8.26 on day 31, after which it gradually increased, stabilizing at 8.41 after day 52. In contrast, group A peaked on day 5, dropped to a low of 8.15 by day 11, then fluctuated upward, ultimately stabilizing around 8.38. The addition of cow manure effectively mitigated the drastic changes in pH during the early to mid-stages of composting, providing a favorable environment for microorganisms and promoting OM decomposition and straw maturation.

The EC indicates the concentration of soluble salts and plays an essential role in assessing the potential phytotoxicity of the final compost product. Appropriate EC levels can meet the requirements for plant growth; however, when EC exceeds 4000 µS/cm, it can severely inhibit seed germination rates across various plant species [32,33]. Figure 1d shows that the EC value in experimental group B exhibited a similar trend to that of the pH, initially rising to a peak of 1933 µS/cm on day two, followed by a rapid decline and a slight increase, resulting in a final EC value of 999.3 µS/cm. In contrast, control group A displayed a fluctuating downward trend, with the EC decreasing from 1645.5 µS/cm at the beginning of composting to a final value of 880.8 µS/cm. This suggests that microbial activity in experimental group B was more vigorous in degrading organic matter, releasing a complex mixture of small organic acids, ammonium salts, phosphates, and other compounds, potentially enhancing the nutrient content.

3.2. Changes in Representative Carbon and Nitrogen Compounds

Carbon and nitrogen are essential nutrients for evaluating fertilizer effectiveness, and their patterns of change can be used to assess the compost maturation process [34]. The OM content in both groups decreased continuously (Figure 2a), owing to the growth and reproduction of microorganisms via decomposition of OM. The degradation of OM in group B was higher than in group A during the initial stage (0–17 days), while the opposite trend was observed in the later stage (31–66 days). The OM content of group B decreased from 23.51% to 18.37% while in group A the reduction was from 23.07% to 18.24%, with corresponding OM degradation rates of 21.86% and 20.94%, respectively. This indicates that the addition of cow manure can enhance the degradation of OM in mesophilic aerobic composting.

Due to changes in temperature and substrate conditions, nitrogen cycling during composting is governed by different reactions at various stages. Among these, NH₃ release is the primary form of nitrogen loss [11]. Figure 2b illustrates the NH₃ emission rate pattern during aerobic composting. In both control group A and experimental group B, NH₃ emission rates rose sharply in the initial composting stage, reaching 47.57 mg/(m²·h) and 37.07 mg/(m²·h) by day 24, respectively. After a steep decline, the rates then rapidly peaked at 56.99 mg/(m²·h) and 53.17 mg/(m²·h). Overall, although the average NH₃ emission rate in group B was slightly lower than in group A, there was no significant difference between the two systems, indicating that the addition of cow manure did not effectively inhibit NH₃ volatilization. As shown in Figure 2b, the NH₃-N content rose to a maximum value of 0.16 mg/g for group B and 0.14 mg/g for group A on day 5, which was attributed to the ammonification of nitrogenous compounds during the initial stage of composting [35]. Notably, from day 5 until the end of composting, the NH₄⁺-N content in group B remained generally lower than in group A (except for the sample at day 52). Conversely, the TN concentration in group B was higher than in group A during both the early and late composting stages (Figure 2c), likely due to the continuous conversion of NH₄⁺-N to NO₃⁻-N by nitrifying microorganisms. However, the TN content in both systems exhibited a trend of initially rising, followed by a decline, and then an increase, resulting in a higher TN content of 8.95 mg/g in group B compared to 8.40 mg/g in group A on day 66. These results suggest that the addition of cow manure may have accelerated the conversion of NH₄⁺-N to NO₃⁻-N while inhibiting the denitrification process, ultimately improving the TN content in the final compost products. In fact, the high nitrogen content in cow manure provides a significant substrate for nitrifying bacteria, such as species of Nitrosomonas and Nitrobacter, enhancing their activity and proliferation. This microbial enhancement accelerates the nitrification process, leading to a more rapid conversion of NH₄⁺-N to NO₃⁻-N [36]. The aerobic conditions maintained during mesophilic composting, combined with the addition of cow manure, create an environment less favorable to the anaerobic bacteria responsible for denitrification. Moreover, the presence of certain microbial communities in cow manure can outcompete denitrifiers for available substrates, thereby suppressing their activity [19]. This competitive inhibition, along with sustained aerobic conditions, may effectively reduce the denitrification process during composting.

Figure 2d illustrates the changes in TOC during the aerobic composting process. Both experimental group B and control group A exhibited a fluctuating downward trend in TOC. At the end of the composting period, the TOC of group B was 105.37 ± 2.24 mg/g, while the TOC of group A was 83.67 ± 26.25 mg/g. The rapid growth and reproduction of microorganisms significantly depleted the organic carbon components within the composting system, while a portion of the organic carbon was extensively converted to CO₂ and released into the atmosphere. This process resulted in a continuous reduction in TOC in the material. Clearly, by the end of the composting process, the TOC loss in group A reached 61.29%, while group B experienced only a 53.32% reduction. This indicates that the addition of cow manure effectively retained the biodegradable organic carbon within the straw composting system. In accordance with the changes in TOC content, both the C/N ratio and T-values during the composting process also exhibited an overall decreasing trend (Figure 2e,f). The C/N ratio and T-values reflect the biological activity and stability of compost products, making them key indicators for estimating compost maturity [37]. In the later stages of composting, the C/N ratio of group B remained consistently lower than that of group A. Ultimately, the C/N ratios for groups A and B were 9.74 ± 1.69 and 9.19 ± 0.32, respectively, both of which meet the maturity standard of less than 15 [38]. Figure 2f shows that the T-value of group B decreased to 0.34 on day 17, then gradually increased to 0.51 before falling to 0.36. In contrast, the T-value of group A peaked at 0.85 on day 11, subsequently declined, and stabilized around 0.5. Overall, during the majority of the composting period, the C/N ratio and T-value of the B composting system were consistently lower than those of system A. This indicates that the addition of cattle manure facilitated the maturation of straw, resulting in a higher quality of the final compost product.

3.3. Heavy Metal Content and Seed Germination Index

As shown in Figure 3a–e, the concentrations of Hg, Pb, As, Cd, and Cr in the thermophilic aerobic composting materials increased with extended composting time. However, these levels still comply with the limits established by the organic fertilizer standards (NY 525-2012) [39] set forth by the Ministry of Agriculture of the People’s Republic of China. At the end of the composting period, the concentrations of As, Cd, Pb, Cr, and Hg in the control group A increased by 14.95%, 19.07%, 32.50%, 6.62%, and 6.38%, respectively, compared to the beginning. In the experimental group B, the levels of As, Pb, Cr, and Hg rose significantly by 51.37%, 47.40%, 7.89%, and an extraordinary 882.5%, respectively. These results indicate that the addition of cow manure to the straw composting system resulted in the release of higher concentrations of heavy metals, including As, Pb, Cr, and Hg, with a particularly significant increase in Hg levels (p < 0.05). In contrast, the increase in Cd levels during the later stages of composting in group A was significantly higher than in group B. Notably, the Cd concentration in group B was the only heavy metal that decreased compared to its initial levels. Overall, as composting time progressed, the heavy metal concentrations in both systems increased to varying degrees. The addition of cow manure further elevated these levels (with the exception of Cd). This increase can be attributed to the decomposition and volatilization of OM during composting, which concentrates the compost volume, as well as the presence of heavy metals inherent in the added cow manure. For Cd, the initial increase in concentration can be attributed to the reduction in compost weight and volume. However, the observed decrease in Cd concentrations during the later stages of composting is likely primarily due to microbial processes, including biosorption [40], bioaccumulation [41], and biotransformation [42], which sequester and stabilize Cd within the compost matrix. These mechanisms, along with the formation of humic substances, effectively reduce the mobility and toxicity of Cd, thereby improving the overall quality and safety of the compost product.

The GI is a widely used parameter to assess the toxicity levels of final compost on plants, as it directly impacts seed germination and seedling growth. It provides insight into the potential phytotoxic effects of compost, influencing agricultural practices and plant health. Previous studies have proved that the sample is essentially nonphytotoxic if the GI ranges from 65% to 100%. Compost product is mature and stable enough to be used as fertilizer when the GI is higher than 100% [43]. The GI of Chinese cabbage seeds throughout the composting process, as shown in Figure 3f, exhibited a trend similar to that of amaranth seeds, generally displaying an upward trend. This indicates that the composting process positively influenced seed germination rates, suggesting improved compost quality over time. For amaranth seeds, the GI significantly decreased in both group A and group B during the early composting phase compared to the blank control (p < 0.05). This decline may be attributed to the presence of volatile humic acids and ammonia produced from the degradation of OM in the initial stages of composting, which can inhibit seed germination. During the mid composting phase, the GI values for both groups rapidly increased, reaching 96.47 ± 6.20% for group A and 95.49 ± 0.52% for group B. By the end of the composting process, the GI values further rose to 121.05 ± 3.83% for group A and 109.88 ± 6.58% for group B, indicating that complete maturation was achieved.

3.4. Microbial Community

The structure of microbial community is shown in Figure 4. As for bacteria, Proteobacteria (with average relative abundance of 54.14%), Bacteroidota (13.62%), Firmicutes (7.30%), Actinobacteriota (6.75%), Planctomycetota (4.26%), Chloroflexi (4.13%), Verrucomicrobiota (3.43%), Bdellovibrionota (1.51%), and Patescibacteria (1.24%) were the most dominant phyla during the whole composting process, and they represented 94.61–99.00% of the total bacteria in different samples (Figure 4a). The top six phyla are widely reported with the ability to degrade complex OM such as lignin and cellulose and dominated in compost samples [44,45]. The relative abundance of Bacteroidota in samples B31 (14.82%) and B66 (15.01%) was significantly higher than in A31 (9.19%) and A66 (7.85%), which may be one of the reasons for the superior composting effectiveness observed in group B. Notably, the number of unique OTUs in the bacterial communities across all stages in group B was significantly higher than in group A (Figure S1a), suggesting that the addition of cow manure promotes microbial diversity in mesophilic aerobic composting. As composting progressed, the relative abundance of Firmicutes in both systems decreased rather than increased (group A: 8.82% → 7.85%; group B: 7.95% → 5.21%), contrasting with previously reported trends in thermophilic aerobic composting [44]. The primary reason for this discrepancy is that the mesophilic aerobic composting technique used in this study maintained temperatures around 25 °C throughout the composting period, much lower than the thermophilic aerobic composting temperatures (>55 °C). Therefore, Firmicutes did not need to produce endospores to withstand high temperatures [45]. Additionally, the abundance of Actinobacteriota, Planctomycetota, and Patescibacteria increased over time in both systems, consistent with patterns observed in previous studies involving sewage sludge composting [46]. This suggests that these bacteria may play a key role in the degradation of small organic molecules during the middle and later stages of composting.

At the genus level (Figure 4b), an obvious succession of bacterial community occurred. Pseudomonas, a typical denitrifying microorganism widely distributed in sewage treatment systems [47], showed the highest relative abundance (12.13%) in the day 0 samples of group A, while other genera were more evenly distributed. However, it sharply decreased during the middle to late stages, with only 1.08% (A66) and 1.22% (B66) remaining in both systems by day 66. This is likely because, in the early stages of composting, the nitrogen source is abundant, promoting the enrichment of denitrifying bacteria. However, as nitrogen is lost through denitrification and ammonia emissions, the abundance of these functional microbes declines sharply. Similar to the trend observed in Pseudomonas, the genera Devosia, Brevundimonas, unclassified_o__Enterobacterales, Sphingobacterium, and Acinetobacter also followed a similar pattern. Brevundimonas, in particular, has been reported as a key microorganism influencing the conversion of NO₃⁻-N and NO₂⁻-N in cattle manure composting [48]. In addition, unclassified_o__Enterobacterales plays an essential role in OM degradation, which enhances soil structure and fertility [49], while Sphingobacterium significantly contributes to DNA degradation by producing enzymes such as DNase, thereby influencing nutrient cycling within the composting process [50]. Acinetobacter is also a commonly observed microorganism in aerobic composting of agricultural waste, often showing high abundance in the early composting stages [51], which aligns with the findings of this study. In contrast to the microbial groups mentioned above, Rhizobium and bacteria specializing in the degradation of recalcitrant organic compounds emerged as the dominant genera from the early stages to the end of the composting process. Bacillus species have been reported to facilitate the breakdown of cellulose and pectin, accelerating the decomposition of OM during the composting process [52]. Qiu et al. [53] also identified certain Bacillus species with capabilities for both nitrification and nitrogen fixation. Unclassified_f__Rhizobiaceae and Allorhizobium-Neorhizobium-Pararhizobium-Rhizobium are typical nitrogen-fixing rhizobia [54]. Their increased abundance during mesophilic aerobic composting supports nitrogen retention in the organic fertilizer. Notably, Pseudomonas showed higher relative abundance in B31 (3.21%) and B66 (1.22%) compared to A31 (2.92%) and A66 (1.08%). Conversely, Allorhizobium-Neorhizobium-Pararhizobium-Rhizobium was more abundant in A31 (2.33%) and A66 (3.49%) than in B31 (1.92%) and B66 (1.66%). This suggests that the addition of cow manure promoted nitrogen transformation but was less favorable for nitrogen retention. Moreover, unclassified_f__Microbacteriaceae (1.56%), Sphingomonas (1.33%), Asticcacaulis (1.24%), and Chitinophaga (0.97%) were also reported to have the ability to degrade macromolecular organic substances (e.g., protein and lignocellulose) into micromolecular humus precursors [55,56,57,58]. Therefore, the bacterial community formed an efficient macromolecular degradation system.

In both systems, only six fungal phyla were identified, with Ascomycota (constituting 67.22% in group A and 48.67% in group B), Basidiomycota (15.71% in group A and 7.49% in group B), and Mucoromycota (1.13% in group A and 0.50% in group B) as the dominant groups in the community (Figure 4c). Ascomycota remained the dominant phylum throughout the composting process, reaching peak abundance during the mid stage, with 97.16% in group A (A31) and 55.65% in group B (B31). Ascomycota is frequently reported as the dominant fungal phylum in composting systems [44]. Certain groups within this phylum have the ability to secrete various enzymes that facilitate the degradation of organic materials, including cellulose and hemicellulose [59]. Furthermore, the relative abundance of Ascomycota and Basidiomycota in group A was significantly higher than that in group B during the later stages of composting (p < 0.05). Additionally, only a small presence of Mortierellomycota and Chytridiomycota was observed in group B during the maturation stage.

At the genus level, unclassified_k__Fungi (accounting for 29.00%), unclassified_o__Sordariales (11.86%), and unclassified_f__Chaetomiaceae (8.67%) were dominant genera, following by Neocosmospora (4.57%), Schizothecium (4.31%), Aspergillus (4.18%), Kiflimonium (3.31%), Rhodotorula (3.04%), Chaetomium (2.83%), Coprinopsis (2.82%), Stachybotrys (2.33%), and unclassified_c__Sordariomycetes (2.07%) (Figure 4d). Many of these genera are closely linked to the degradation of complex macromolecules, such as lignocellulose. At the initial stage (0 days), Aspergillus, a typical thermophilic fungus known for its OM degradation capabilities [60], exhibited the highest abundance in both groups, accounting for 9.45% in group A and 15.16% in group B. However, it is important to note that several Aspergillus species are opportunistic pathogens [61]. During the mid to late stages of composting, the abundance of Aspergillus in group B was lower than in group A. This finding indicates that the addition of manure promotes the degradation of straw in the early composting phase and reduces pathogen levels in the mid to late stages. This is consistent with the OTU findings, where group B showed a higher number of unique OTUs than group A only during the initial composting stages (Figure S1b). Similar phenomena were found in Neocosmospora and Fusarium, which were also pathogenic bacteria to plants or animals [62]. The reduction in these pathogens through the addition of fresh manure in this study enhances the potential for safely using compost products in subsequent applications.

Initially, the distribution of dominant fungal genera was relatively uniform. However, from day 31 onward, unclassified_o__Sordariales emerged as the most abundant genus, comprising 28.84% in group A and 8.17% in group B. This genus persisted until the end of the composting process, with proportions of 6.80% in group A and 27.32% in group B, indicating a clear shift in community structure. Certain genera within Sordariales, such as Mycothermus and Thermomyces, have been reported to play significant roles in cattle manure composting by producing cellulases and hemicellulases [63]. However, their effectiveness may be more pronounced in the later stages of the composting process in this study.

3.5. Metabolism of Microorganisms

To investigate biomass degradation and carbon/nitrogen metabolism, we performed functional predictions of the microbial communities in each sample. Figure 5a illustrates the abundance of pathways related to “Degradation/Utilization/Assimilation” in fungi, as derived from the MetaCyc database. The degradation pathways of amino acids, sugars, and fatty acids, as well as the N, P, and S metabolism in fungi, exhibited high abundance during the mid to late stage of composting, which is consistent with the abundance of dominant fungal genera (Figure 4d). Additionally, in the early stage of the mesophilic aerobic composting process, the abundances of these pathways in group B were higher than in group A. These results suggest that fungi predominantly contributed to the later stages of composting, while the addition of cow manure enhanced the degradation of OM in the early stages by stimulating indigenous fungal activity. This phenomenon is in contrast to traditional thermophilic aerobic composting [44], indicating that the microbial succession patterns and underlying mechanisms in mesophilic aerobic composting still require further investigation.

The bacterial pathways, based on the KEGG and MetaCyc databases, are shown in Figure 5b and Figure 5c, respectively. In contrast to fungi, the abundance of carbon and nitrogen metabolism pathways in bacteria was higher during the early stages of composting than in the mid to late stages (Figure 5b). This suggests that bacteria primarily play a role in the initial phase of mesophilic aerobic composting, when carbon, nitrogen, phosphorus, and sulfur are more abundant. Additionally, during the early stages of mesophilic aerobic composting, group B exhibited higher abundance of these pathways compared to group A, further confirming that the addition of cow manure enhanced biomass decomposition and carbon/nitrogen metabolism, particularly amino acid and carbohydrate metabolism. The incomplete degradation in group A during the early stages likely explains the higher abundance of these pathways in the later stages of composting (Figure 5b). Amino acid metabolism provides essential energy and carbon sources for bacterial growth and reproduction. It has also been reported that amino acids are key precursors for humus formation during the humification process in composting [64]. Therefore, the increased abundance of amino acid metabolism in group B facilitated the synthesis of humic substances. However, denitrification exhibited a higher abundance in group B (Figure 5c), suggesting that the addition of cow manure may have increased nitrogen loss by enhancing the denitrification process.

To further investigate nitrogen metabolism during composting, the abundance of functional genes associated with nitrogen metabolism in bacteria was analyzed based on the predicted results (Figure 5d). Proteases are essential enzymes in the ammonification process, with alkaline metallopeptidase (encoded by apr) and neutral metallopeptidase (encoded by npr) being the most significant. These enzymes contribute to 20% and 70% of the total extracellular protease activity, respectively [65]. The abundances of these genes remained relatively high throughout the composting process, from the early stages to the end. This suggests extensive protein degradation and a significant release of NH₃ during this period, which aligns with the observed trends in NH₃-N content and NH₃ emissions (Figure 2b).

Traditional nitrification consists of two steps: the ammonium oxidation process (NH₄⁺-N→NO₂⁻-N) and the nitrite oxidation process (NO₂⁻-N→NO₃⁻-N). In terms of the nitrification process, the genes amoA, amoB, and amoC exhibited the highest abundance during the middle and later stages of composting in group A, with the overall abundance in group A being higher than that in group B. The abundance of the hydroxylamine oxidase gene (hao) was also higher in the later stages of composting compared to other stages. However, in contrast to the ammonia monooxygenase genes, the hao gene was significantly more abundant in group B than in group A. These results suggest that while the addition of cow manure inhibited the ammonium oxidation process in the medium-temperature aerobic composting, it promoted the nitrite oxidation process, leading to a greater conversion of NH₄⁺-N to NO₃⁻-N rather than to NO₂⁻-N. It should be noted that although the nitrite oxidation process was somewhat enhanced, the ammonium oxidation process was significantly inhibited. Therefore, overall, nitrification in the cow manure-added group remains suppressed.

In the denitrification process, the abundance of nitrate reductase-encoding genes (narG, narH, narI, napA, and napB) was lower in the later stages of composting compared to earlier stages. Moreover, the abundances of these genes in group B were consistently higher than those in group A throughout all stages of composting. This pattern was also observed for the genes encoding nitrite reductase (nirK and nirS) and nitrous oxide reductase (norZ). In contrast to the trends observed for the nitrate reductase- and nitrous oxide reductase-encoding genes, the abundance of nitric oxide reductase-encoding genes (norB and norC) in group B was higher than in group A during the early stages of composting. It is inferred that the addition of cow manure may have enhanced nitrogen loss by inhibiting nitrification and promoting the denitrification process within the microbial community. As a result, nitrogen was lost from the piles in the form of N₂ and N₂O. Furthermore, no genes related to anammox (hzs and hdh) were detected in any of the samples, which is consistent with the findings of Tian et al. [44] in their study on sewage sludge composting.

The assimilation process involves the assimilatory reduction of NO₃⁻-N and NO₂⁻-N, as well as the incorporation of NH₃-N into organic nitrogen. In this study, the abundance of narB was highest in the early stages of group B, while nasC, nasB, and nirA exhibited the highest abundance in the later stages of group A. Notably, the genes involved in assimilatory nitrate reduction (narB, nasC, and nasB) were more abundant in the early stages of group B compared to group A. As for the core pathways of ammonium assimilation, the gene abundance of glutamine synthetase (glnA) was highest in the later stages of group A, while it exhibited a higher abundance in the early stages of group B. It was speculated that the addition of cow manure might promote the assimilation processes of nitrate and nitrite during the early stages of the mesophilic aerobic composting. Additionally, the abundance of genes associated with nitrogen fixation (nifD, nifK, nifH, and nifG) in group A was significantly higher than that in group B during the later stages of composting. In conjunction with the abundance of nitrogen-fixing bacteria (Figure 4b), these findings suggest that the addition of cow manure may suppress the activity of most nitrogen-fixing bacteria.

Composting involves a succession of microbial activities that transform organic nitrogen into inorganic forms, primarily through ammonification, nitrification, and denitrification. The temperature regimes in mesophilic (moderate temperature) and thermophilic (high temperature) composting significantly influence these nitrogen transformation pathways. But the nitrogen transformation mechanisms in mesophilic and thermophilic composting differ notably due to the temperature-dependent microbial activities. Mesophilic composting offers a more controlled nitrogen transformation with potentially lower nitrogen losses but at a slower rate. In contrast, thermophilic composting accelerates organic matter decomposition and nitrogen transformation but may incur higher nitrogen losses due to volatilization and denitrification.

3.6. Correlation Analysis of Environmental Factors, and Microorganisms

Network analysis further revealed the relationships between microbes and environmental factors (Figure 6a–d). These graphs depict only nodes with significant correlations, defined by a Spearman correlation coefficient of r > 0.5 and p < 0.05. Overall, the complexity of major bacterial–environmental factor interactions in group A was significantly higher than in group B, suggesting that the addition of cow manure reduced the structural and community complexity of bacteria in the mesophilic aerobic composting system. In the correlation network of dominant bacteria and environmental factors in group A, the dominant bacteria showed positive correlations with NH₄⁺-N, TN, and the C/N ratio, while displaying negative correlations with pH and NH₃ (Figure 6a). This finding indicates that these bacteria played a significant role in breaking down OM, thereby accelerating the decomposition of straw material. In the cow manure addition group, however, most bacteria displayed a significant negative correlation with pH, temperature, TOC, TN, and NH₃, while exhibiting a positive correlation with EC and the C/N ratio (Figure 6b). Pseudomonas and Brevundimonas were positively correlated with the C/N ratio and NH₄⁺-N, reflecting their roles in OM degradation. This association aligns with their observed gradual decline in abundance during the middle and later stages of composting (Figure 4c). Additionally, it is important to note that Pseudomonas, Brevundimonas, and norank_f__Microscillaceae showed positive correlations with the C/N ratio, NH₃, and NH₄⁺-N. This suggests that these genera might be involved in the nitrification process, potentially contributing to nitrogen loss during composting. Additionally, moisture content has been reported to positively correlate with N₂O emissions [66]. In this study, only Asticcacaulis showed a positive correlation with moisture content, suggesting that this microbe may have the potential to promote N₂O emissions.

The network analysis of dominant fungi and environmental factors in group A showed that Aspergillus, Alternaria, and Rhodotorula had positive correlations with NH₃-N, TN, and the C/N ratio. This suggests that, without the addition of cow manure, these fungi played a vital role in the breakdown of wheat straw OM (Figure 6c). Additionally, Aspergillus, Alternaria, Rhodotorula, and Mortierella exhibited negative correlations with NH₃ and pH, which may indicate that these genera also participated in the nitrification process, thereby promoting nitrogen retention. The network analysis of dominant fungi and environmental factors in group B showed a more complex relationship (Figure 6d), indicating that the addition of cow manure enhanced fungal growth and metabolism. This, in turn, facilitated the more pronounced role of fungi in the mesophilic aerobic composting system. In group B, only Dipodascaceae_gen_Incertae_sedis and Rhodotorula showed a positive correlation with NH₃-N, reflecting their involvement in OM degradation. Furthermore, Irpicaceae_gen_Incertae_sedis was positively correlated with temperature, TN, and TOC, but negatively correlated with NH₃, suggesting its potential role in different stages of composting and its involvement in nitrogen dynamics. Aspergillus, Neocosmospora, Rhodotorula, Solicoccozyma, Cladosporium, and Dipodascaceae_gen_Incertae_sedis were also negatively correlated with NH₃, suggesting that these genera may be involved in the nitrification process and could contribute to nitrogen retention in the composting system.

4. Conclusions

The addition of cow manure in mesophilic aerobic composting significantly influenced nitrogen transformations, microbial succession, and compost stability. Group B, enriched with cow manure, demonstrated enhanced ammonification, nitrification, and accelerated early decomposition of organic matter, contributing to improved nitrogen retention and expedited compost maturation. Microbial network analysis revealed that cow manure addition simplified bacterial–environment interactions while enhancing fungal activity, thereby promoting efficient organic degradation. The final compost products met maturity standards, making mesophilic aerobic composting with cow manure a viable approach for managing agricultural waste and optimizing nitrogen retention. These findings contribute to advancing sustainable composting methodologies and highlight the need for further research into nitrogen retention strategies for mesophilic composting systems.