1 Introduction

Aerobic composting is a process in which microorganisms decompose organic waste, converting it into humus, which can be directly used as fertilizer for the fields (Qian et al. 2014). However, this composting process generates a large amount of volatile gas, primarily NH3, due to biochemical reactions. The volatilization of ammonia not only results in significant nitrogen (N) loss but also contributes to air pollution. Studies indicate that approximately 60% of N loss during aerobic composting is due to NH3 emissions (Yang et al. 2020). Thus, controlling NH3 emissions during composting is crucial for reducing environmental risks and improving compost quality.

NH3 emissions during composting are mainly influenced by factors such as temperature (Kianirad et al. 2010), pH (Cao et al. 2020), moisture content (Ding et al. 2019), aeration rate (Zhang et al. 2013), and C/N ratio (Jiang et al. 2011). Adjusting composting materials, environmental conditions, and process parameters can be costly and challenging. Presently, using conditioners offers clear advantages in reducing N loss during composting. Conditioners come in various forms, including biological, physical, and chemical conditioners (Shan et al. 2021). Biological conditioning involves adding microbial agents during composting, such as high-temperature ammonia-resistant (TA T105H) (Kuroda et al. 2015) and ammonia-oxidizing bacteria (AOB) (Zhang et al. 2016). These microbial inoculants reduce NH3 emissions by promoting NH4+ assimilation or nitrification and lowering NH4+ levels (Cao et al. 2019). However, the choice of suitable microbial inoculants depends on the composting materials and parameters.

Physical conditioning involves adding materials with good pore structure, like biochar and zeolite, to adsorb and immobilize NH3. Biochar is a popular physical conditioner due to its large pore structure and abundant surface functional groups (Liu et al. 2017). Yet, its ability to continuously adsorb NH3 may be affected by the saturation of adsorption sites. Chemical conditioning focuses on adjusting the pile's pH and the conversion of carbon and nitrogen by adding additives like superphosphate and phosphogypsum. Chemical conditioning is more effective than biological or physical conditioning in reducing NH3 emissions (Cao et al. 2019). However, the use of chemical conditioners may lead to increased reactor conductivity, incomplete decomposition, and inhibition of seed germination (Wang et al. 2020).

The effect of iron conditioners on composting has garnered significant attention in recent years. Several studies have demonstrated the effectiveness of iron conditioners in reducing NH3 emissions during composting. For instance, Liu et al. (2022) found that 3% zero-valent iron (ZVI) effectively reduced NH3 cumulative emissions by 9.4% in sludge composting. Similarly, Chen et al. (2022) reported that the addition of 5% FeSO4 reduced NH3 emissions by 12.4% during pig manure composting. Despite their efficacy, these iron conditioners have limited practical applications due to their poor stability. FeCl3, a common acidic salt material with good stability, presents a potential alternative for reducing NH3 emissions during composting. In recent years, researchers have rarely studied ferric chloride as an aerobic regulator. And little is known about the changes in physicochemical properties and the effect of NH3 emissions in composting with the addition of FeCl3. Therefore, it is necessary to investigate the effect of adding FeCl3 on the composting process. However, adding FeCl3 alone may result in a high concentration of salt ions in the compost, which can impact material maturation and adversely affect seed germination. To address this issue, the addition of biochar conditioner, which possesses a buffering effect and reduces phytotoxicity in composting, could prove beneficial. Combining FeCl3 and biochar may be more effective in reducing NH3 emissions. In recent years, the use of FeCl3 and the combination of FeCl3 and biochar as a compost conditioner has not been reported yet. The mechanisms of their effects on the physicochemical characteristics and NH3 emissions of the composting process deserve to be explored by researchers.

In this study, the researchers hypothesize that co-conditioning aerobic composting with Fe/BC could lead to complementarity between the two additives, resulting in significantly reduced NH3 emissions and improved organic fertilizer quality. The objectives of the study were (a) to investigate the effects of FeCl3 and biochar on NH3 emissions during aerobic composting and analyze the potential mechanism behind emissions reduction; and (b) to examine the physical and chemical properties, nutrient dynamics, organic matter, and humus content of the Fe/BC composting process and further evaluate its overall impact on the quality of organic fertilizer. Through this research, the study aims to present a new approach for reducing the risk of NH3 emissions and enhancing the quality of organic fertilizers in the composting process by utilizing Fe/BC co-conditioners.

2 Materials and methods

2.1 Experimental materials

The main ingredients used in aerobic composting were chicken manure and wheat straw, both of which were collected from a local farm in Nanjing, Jiangsu Province. The wheat straw was cut into small pieces measuring 1–2 cm for better utilization. Additionally, wood biochar was utilized in the composting process, and its preparation involved a temperature of 400 °C, resulting in a biochar with a pH of 7.82, surface area of 6.84 m2 g−1, total pore volume of 0.018 cc g−1, and an average pore diameter of 8.98 nm. For further details on the properties of the composting materials, please refer to Additional file 1: Table S1 and Fig. S1.

The composting reactor used in this study was comprised of incubators with an outer diameter measuring 64 × 48 × 36 cm. The case’s shell and inner liner were constructed from PP material, and the center featured a 5 cm-thick polystyrene foam, providing thermal insulation functionality. To monitor the temperature during composting, an RC-4 automatic thermometer was utilized.

For further analysis of the composting materials, the FI-IR spectra and 2D-COS analysis can be referred to in Additional file 1: Figs. S3 and S4, respectively. These additional figures provide valuable information on the composting process and its materials.

2.2 Treatments

The aerobic composting experiment took place in a greenhouse located at the Jiangsu Academy of Agricultural Sciences (118.87° E, 32.03° N). For the experiment, chicken manure and wheat straw were thoroughly and evenly mixed at a dry weight ratio of 1.5:1. Each composting reactor’s initial dry weight of organic waste was 8 kg, with an initial C/N ratio of approximately 20–25, and the initial moisture content was maintained at around 60%. The entire composting experiment spanned a period of 30 d.

To investigate the impact of different additives on the composting process, five distinct treatment groups were established based on different additive ratios (Control group without any conditioner application (CK); F1: Composting with 1% FeCl3 (w/w) additive; F3: Composting with 3% FeCl3 (w/w) additive; F1 + B15: Composting with a combination of 1% FeCl3 and 15% biochar (BC, w/w) as additives; F3 + B15: Composting with a combination of 3% FeCl3 and 15% biochar (BC, w/w) as additives).

2.3 Sample collection and analysis

2.3.1 Sample collection

The composting samples were subjected to full turning on the 4th, 7th, 10th, 13th, 17th, and 21st day. A total of 200 g of samples were collected from the center and four corners of the compost pile and mixed thoroughly to ensure homogeneity. Subsequently, 50 g of these samples were transported back to the laboratory for testing, while the remaining samples were returned to the composting reactor.

Each collected sample was divided into two parts. One part was frozen and stored for later analysis of physical and chemical properties, while the other part was dried and crushed to determine nutrient content, organic matter, and humus levels.

2.3.2 NH 3 collection and testing

NH3 was collected using a closed batch pumping technique with boric acid absorption. The collection setup comprised a sealed plexiglass cylinder with a 10 cm diameter, featuring air inlet and air outlet at the top. The inlet hole was connected to a vent pipe, which extended to a height of 2 m. Additionally, the outlet hole was linked to a vacuum pump through the absorption device. During the composting cycle, NH3 was collected for a duration of 4 h each day. The collected NH3 was then titrated using 0.02 mol L−1 sulfuric acid, and 2% boric acid was utilized as the absorbent solution. Furthermore, the titration process employed methyl red-bromocresol green as an indicator to detect the endpoint.

2.3.3 Physicochemical index determination

Temperature recording during composting involved inserting a temperature recorder into the center of the compost pile at the beginning of the process. The temperature readings were taken every half hour, throughout the day. To analyze the data, the average temperature of the day was calculated, and the ambient temperature was also measured. Additionally, this data was used to assess the feasibility of collecting ammonia volatilization data in the system. The effective accumulated temperature equation is as follows:

$$T = \Sigma \left( {T_{i} - T_{0} } \right) \times \Delta t$$
(1)

where Ti is the pile temperature at the moment i, T0 is the initial temperature (biological zero) at the time of microbial proliferation in the compost, and Δt is the duration of Ti.

Determination of pH, EC, and GI: Fresh composting samples were mixed with deionized water at a ratio of 1:10 (solid–liquid ratio). The mixture was then oscillated for 30 min to ensure proper mixing. After oscillation, the mixture was left to stand for 10 min to allow any solid particles to settle. Subsequently, the supernatant was collected for further testing. The collected liquid sample was used to measure the pH and EC. This was accomplished using a multi-parameter analyzer. According to Shan et al. (2019), the GI (Germination Index) was measured and calculated using the following formula:

$$GI \left( \% \right) = \frac{{Seed\;germianation{ }\left( \% \right)*root\;length\;of\;treatment}}{{Seed\;germianation{ }\left( \% \right)*root\;length\;of\;control}} *100\%$$
(2)

2.3.4 Nutrient determination

Determination of NH4+–N and NO3−–N: Fresh compost samples and 2 mol L−1 KCl solution were mixed in the ratio of 1:10 (w/v). The samples were shaken for 30 min and allowed to stand for 10 min for filtration. The filtrate was analyzed using a flow analyzer, and a total of 3 replicates were performed.

Determination of TN, TP, and TK: The H2SO4-H2O2 method was employed to process the composting samples. Subsequently, the total nitrogen content of the processed samples was determined using a Kjeldahl nitrometer. Additionally, the total phosphorus (P) and potassium (K) levels were analyzed using inductively coupled plasma optical emission spectroscopy (ICP-OES, Optima TM8000, Perkin Elmer) (Bartos et al. 2014).

2.3.5 Determination of organic matter and humus

Determination of Humus (HS) Content: A 2.50 g air-dried compost sample was taken and added to sodium hydroxide and disodium hydrogen phosphate extract. The resulting supernatant was collected to determine total organic carbon (TOC), representing the total humus content. From this, 20.00 mL of the supernatant was taken and 6 mol L−1 HCl was added to precipitate the solution. The resulting supernatant contained fulvic acid (FA), while the precipitated substance was humic acid (HA). The content of fulvic acid (FA) was determined using a TOC instrument, and the humic acid content was obtained by calculating the difference between the total humus content and the fulvic acid content (Wang et al. 2016).

Determination of Organic Matter (OM): The organic matter (OM) content of the composting material was determined using the potassium dichromate volumetric method. Additionally, the dissolved organic carbon (DOC) was measured using an organic carbon analyzer (Wang et al. 2016).

Determination of Dissolved Organic Matter (DOM): To analyze the different types of DOM, 5 g of compost samples were dissolved in 50 mL of deionized water (w/v) and then subjected to centrifugation for filtration (Usman et al. 2019). The 3D-EEM technique was employed with emission wavelengths (Em) and excitation wavelengths (Ex) ranging from 200 to 900 nm. Three parameters of the fluorescence spectrum were analyzed: fluorescence index (FI), biological index (BIX), and humification index (HIX).

2.3.6 Quantitative PCR of AOA-amoA and AOB-amoA gene

DNA Extraction and Quantification: DNA was extracted from a 0.25 g compost sample and quantified using fluorescence quantitative PCR (ABI7500, Applied Biosystems, USA). For the AOA-amoA gene, the upstream primer was “STAATGGTCTGGGCTTAGACG,” and the downstream primer was “GCGGCCATCCATCTGTATGT.” For the AOB-amoA gene, the upstream primer was “GGGGTTTCTACTGGTGGGT,” and the downstream primer was “CCCCTCKGSAAAGCCTTCTTC.”

qPCR Amplification Conditions: The qPCR amplification was conducted with the following conditions:

  • 1. Pre-denaturation at 94 °C for 15 min.

  • 2. Cycles: a. Denaturation at 94 °C for 30 s, b. Annealing at 55 °C for 30 s, c. Extension at 55 °C for 30 s, d. Final extension at 55 °C for 30 s.

  • 3. Extension at 72 °C for 10 min.

Quantification of Copy Number: The copy number of each DNA sample was determined using the amplification standard curve method.

2.3.7 Statistical analysis

Microsoft Excel (2022) was utilized for data processing, while SPSS 27 was employed for conducting statistical analysis. Univariate analysis of variance (ANOVA) was used to calculate the significant differences (P < 0.05). The image was created using Origin (2023b). The 3D-EEM images were generated using the R language. The models were validated through disassembly analysis, model fitting, and visual inspection of residuals.

3 Results and discussion

3.1 NH 3 emissions during Fe/BC co-conditioning composting

Figure 1a illustrates the daily curve of NH3 emissions during composting. The composting process encompassed three distinct stages: high-temperature period (HTP, Days 1–7), cooling period (CP, Days 8–19), and maturation period (MP, Days 20–28). Notably, NH3 emissions exhibited a fluctuating pattern in response to reactor turning. During the high-temperature period, microorganisms effectively degraded amino acids, proteins, and other readily degradable organic matter, leading to the production of NH3. The CK treatment experienced its peak NH3 emissions primarily in the high-temperature period of composting, with a peak intensity of 4.10 g kg−1 d−1, consistent with conventional aerobic composting trends (Zhang et al. 2014). On the other hand, for the F1, F3, F1 + B15, and F3 + B15 treatments, the peak NH3 emissions were more pronounced during the cooling period, with the F1 + B15 treatment demonstrating the highest peak intensity (1.43 g kg−1 d−1). Remarkably, these treatments exhibited a significant reduction in peak NH3 emissions (65.1–72.2%) compared to the CK treatment. This outcome aligns with the findings of Chen et al. (2022), who observed a similar reduction in NH3 emissions (43.7%) and TN loss (33.8%) upon the addition of 5% FeSO4. The decrease in NH3 emissions can be attributed to the pH-lowering effect of the Fe conditioner during the composting process, along with the increased NH4+–N content in the high-temperature period, which effectively curtails NH3 emissions.

Fig. 1
figure 1

a Daily ammonia emissions during composting; b cumulative ammonia emissions at the end of composting. HTP High temperature period, CP cooling period, MP maturation period, Turn Turn the composting samples. Data are mean ± SD (n = 4)

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Figure 1b displays the cumulative NH3 emissions during composting. Regarding the NH3 cumulative emissions during the high-temperature period of composting, the CK treatment exhibited the highest NH3 cumulative emissions, while the other treatments showed reductions ranging from 77.3% to 95.7% compared to the CK treatment. In terms of NH3 cumulative emissions during the cooling period of composting, the F1, F3, and F1 + B15 treatments recorded higher NH3 cumulative emissions than the CK treatment. Notably, among these, the F1 treatment had the highest cumulative emissions of NH3, measuring 11.40 g kg−1, accounting for 70.1% of the cumulative emissions of NH3 throughout the treatment cycle. In relation to NH3 cumulative emissions during composting maturity, the F3 treatment exhibited the highest NH3 cumulative emissions, reaching 5.33 g kg−1. Additionally, the cumulative NH3 emissions for each treatment were lower than those during the cooling period.

Considering the cumulative emissions throughout the entire composting cycle, the F1, F3, F1 + B15, and F3 + B15 treatments demonstrated significant reductions ranging from 11.1% to 48.2% in NH3 emissions compared to the CK treatment. The NH3 emissions reduction effect was more pronounced in the F3 treatment compared to the effect observed in the case of the F1 treatment, suggesting a positive correlation between the amount of FeCl3 addition and NH3 emissions. Presumably, higher FeCl3 levels were more effective in reducing pH levels and retaining NH4+. Moreover, the F3 + B15 treatment exhibited the greatest potential in reducing NH3 emissions during composting. Apart from the role of FeCl3 in NH3 emissions reduction, biochar also played a significant role, owing to its unique porous structure and extensive specific surface area, which facilitated the absorption of NH3 released during composting (Xiao et al. 2017). Furthermore, biochar influenced NH4+–NH3 balance and nitrogen mineralization dynamics, thereby contributing to the reduction of NH3 emissions (Awasthi et al. 2017).

3.2 Physicochemical properties during Fe/BC co-conditioning composting

3.2.1 Composting temperature

Temperature is a key parameter in the process of aerobic composting (Greff et al. 2022), and its temperature change will affect the activity of different kinds of microorganisms, thus affecting the decomposition rate of organic matter.

The temperature variation trend of each treatment was basically the same (Fig. 2a). All treatments heated up rapidly on the first day of composting. When the temperature reaches 50 °C, it can be considered to enter the HTP. In this stage, thermophilic microorganisms decompose complex organic matter (cellulose, protein, etc.) in the material, resulting in the continuous rise of the compost temperature (Sudharsan Varma and Kalamdhad 2015). CK, F1, F3, F1 + B15 and F3 + B15 maintained above 50 °C for 6–7 days. The peak temperature of each treatment was 60.0–69.0 °C, and then the temperature began to decrease gradually, and finally the temperature gradually stabilized at 26.2–29.2 °C. The peak temperature of F3 and F3 + B15 treatments in the high temperature stage of composting was significantly lower than that of other treatments, and the time of entering the high temperature stage was significantly delayed. After the first 4 turns (on Day 4, 7, 10, 13, respectively), the temperature of each treatment showed a rapid rise. This is mainly due to the increase of oxygen content in the composting after turning over, which is conducive to the growth and reproduction of microorganisms. The temperature of each treatment had no significant change after the last two turns (on Day 17 and 21 respectively).

Fig. 2
figure 2

a Daily temperature variation during composting; b accumulated temperature of the pile at the end of composting; c the pH during composting; d the EC during composting; e the GI during composting; f the GI at the end of the composting; HTP high temperature period, CP cooling period, MP maturation period. Data are mean ± SD (n = 3)

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Figure 2b shows the accumulated temperature during composting. The effective accumulated temperature of all treatments exceeded 10,000 °C. Compared with CK, the effective accumulated temperature of F1 + B15 and F3 + B15 increased by 1.4–3.3%. In terms of the high temperature period in the composting process, accumulated temperature of F1 and F1 + B15 treatments decreased slightly compared with CK. However, the accumulated temperature of F3 and F3 + B15 decreased significantly by 17.0% and 13.1%. During the cooling period, the accumulated temperature of F1 treatment decreased, while the accumulated temperature of F1 + B15, F3 and F3 + B15 increased significantly compared with CK, reaching 14.0–49.9%.

3.2.2 Composting pH and EC

The pH value is an important parameter in aerobic composting. The pH will affect the activity of microorganisms and indirectly affect the process of aerobic composting (Chan et al. 2016). The optimal pH range for aerobic compost is 6.5–8.5 (Proietti et al. 2016). The pH of each treatment in the composting process firstly increased and then decreased (Fig. 2c). Compared with CK, the pH values of the other treatment groups were lower in the high temperature period. This is because the hydrolysis of FeCl3 will produce a large amount of H+, resulting in a significant decrease in the pH value. This may also be the main reason for reducing NH3 emissions during HTP. However, at the end of composting, the pH of F1 and F3 decreased slightly compared with CK, while pH of F1 + B15 and F3 + B15 increased by 1.5–3.2%, and the difference was significant (P < 0.05). The pH value in the maturation period decreased compared with that in the cooling period, which may be related to CO2 emissions and nitrification process (Liu et al. 2017).

The EC reflects the soluble salt content in the material and indirectly reflects the change of the internal environment of the composting. The EC values for all treatments increased with time from 1.03–1.58 ms cm−1 at the beginning to 3.3–3.98 ms cm−1 at the end of composting (Fig. 2d). This may be due to the continuous release of soluble salts from the decomposition of organic matter in the composting (Gómez-Brandón et al. 2008). The EC values of F3 and F3 + B15 treatments were significantly higher than those of other treatments, which may be due to the fact that the introduction of a high amount of FeCl3 will significantly increase the ion content in the composting. The final EC value of F3 + B15 treatment was lower than that of F3 treatment, which was possibly ascribed to the reduction of salt ions by biochar adsorption and mineral salt precipitation processes (Chung et al. 2021).

3.2.3 Germination index (GI)

It is necessary to determine the toxicity of organic fertilizer after composting to determine whether it can be used in agricultural production. Seed germination is the initial stage of plant growth, which can be sensitive to the external growth environment. Therefore, seed germination test is widely used to evaluate the phytotoxicity of compost products and the degree of compost maturation (Yang et al. 2021).

The GI values of all treatments were lower in the early composting stage (Fig. 2e). This is because microbial decomposition of organic matter will produce more ammonium salts and organic acids, and these substances have an inhibitory effect on seed germination. At the same time, the addition of FeCl3 led to a decrease in the pH of the heap, which might also inhibit the germination and growth of plant seeds. With the consumption of organic acids, pH increased and germination index showed a rapid rising trend.

Compared with CK, the GI values of F1 and F3 treatments was reduced to a certain extent, but still higher than 70% in the end, reaching the harmless standard for composting (Fig. 2f). In contrast, the GI value of F1 + B15 and F3 + B15 was 88.7% and 84.7%, respectively, which was higher than that of CK at the end of composting. This is because the added biochar is alkaline and can act as a buffer against the acidity of FeCl3 hydrolysis. Additionally, the content of heavy metal ions, including As, Cd, Cr, and Pb in the final composts was within the standard specified range (Additional file 1: Table S2).

3.3 Nutrient content during Fe/BC co-conditioning composting

3.3.1 NH 4 + –N and NO 3 − –N content

NH4+–N and NO3−–N represent essential inorganic nitrogen forms involved in nitrogen conversion during composting. As depicted in Fig. 3a, the NH4+-N content in all treatments exhibited an initial increase followed by a subsequent decrease. Upon completion of composting, the F1, F3, F1 + B15, and F3 + B15 treatments displayed significantly higher NH4+-N content than the CK treatment, with the F3 + B15 treatment registering the highest levels. This observation can be attributed to the hydrolysis of FeCl3, which creates an acidic environment facilitating NH3 absorption and consequent NH4+–N content enhancement in the composting process (Cao et al. 2020). Moreover, numerous studies have demonstrated that the addition of biochar to composting enhances NH4+ adsorption due to its remarkable ion adsorption capability.

Fig. 3
figure 3

a NH4+–N and b NO3−–N content during composting; c TN, TP and TK and d total nutrient content at the end of composting; HTP high temperature period, CP cooling period, MP maturation period. Data are mean ± SD (n = 3)

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The content of NO3−–N in all treatments displayed gradually increasing trend, as illustrated in Fig. 3b. After composting, the F1, F3, F1 + B15, and F3 + B15 treatments exhibited an 9.5–27.5% increase in NO3−–N content compared to the CK treatment. The low initial pH of the F1, F3, F1 + B15, and F3 + B15 treatments may have inhibited the activity of nitrifying bacteria, contributing to this pattern (Mei et al. 2021). Additionally, nitrate reduction resulted in the transformation of NO3−–N into NH4+–N, consistent with the changes observed in NH4+–N content. Furthermore, the NO3−–N content in the F1 + B15 and F3 + B15 treatments surpassed that of the F1 and F3 treatments. The inclusion of biochar increased the porosity of the composting materials and improved oxygen levels within the piles, thereby promoting the nitrification process and leading to the accumulation of NO3−–N (Khan et al. 2014).

3.3.2 TN, TP, and TK content

Nitrogen content serves as a critical indicator in composting products. As depicted in Fig. 3c, the TN content underwent significant changes after composting. Comparatively, the TN content in the F1, F3, F1 + B15, and F3 + B15 treatments demonstrated a substantial increase of 27.2–51.1% when compared to the CK treatment (P < 0.05). However, there was no significant difference in TK content. The conversion of nitrogen-containing organic matter into nitrogen oxide, accompanied by NH3 release, played a role in this process. The composting process in F1, F3, F1 + B15, and F3 + B15 treatments experienced an inhibition of microbial activities related to nitrogen decomposition due to the low pH value, leading to a decreased decomposition rate of nitrogen-containing organic materials (Luo et al. 2013). Although TP content in the F1 treatment showed a significant difference compared to CK (P < 0.05), no significant difference was observed in TP content in the F3, F1 + B15, and F3 + B15 treatments. Throughout the composting process, phosphorus exhibited relatively greater stability when compared to nitrogen. Despite various forms of phosphorus converting among each other, the absolute content of total phosphorus remained largely unchanged.

In Fig. 3d, the changes in total nutrients after composting were represented. The total nutrient content of compost was determined by calculating the mass fraction of TN + P2O5 + K2O (dry weight). At the end of composting, there was no significant difference in the total nutrient content of the F1 treatment compared to CK. However, the total nutrient content of the F3, F1 + B15, and F3 + B15 treatments demonstrated an increase of 6.8–15.3%.

3.4 Organic matter and humus content during Fe/BC co-conditioning composting

3.4.1 Organic matter content

During the composting process, microorganisms continuously decompose organic matter, leading to a gradual decline in organic matter content (Awasthi et al. 2016). After composting, the organic matter content in the CK treatment was the highest, reaching 65.8% (Fig. 4a). In comparison, the organic matter content in the F1, F3, F1 + B15, and F3 + B15 treatments exhibited a slight reduction when compared to CK. This finding suggests that the addition of Fe/BC improved the final degree of organic matter degradation, which aligns with results obtained by Sudharsan Varma and Kalamdhad (2015).

Fig. 4
figure 4

a Organic matter and b DOC content at the end of composting; c total humus content at the end of composting; d HA/FA ratio at the end of composting. Data are mean ± SD (n = 3)

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Figure 4b displays the content of DOC at the end of composting. Throughout the composting process, microorganisms mineralize organic carbon to acquire energy for their growth and reproduction (Wong and Fang 2000). The content of DOC in each treatment exhibited significant changes after composting. Compared to the CK treatment, the content of DOC in the F1, F3, F1 + B15, and F3 + B15 treatments underwent a significant reduction of 33.1–62.1%. It is plausible that the Fe/BC co-conditioner enhanced the porosity of the composting piles, thereby promoting microbial activity. Consequently, the readily degradable part of organic matter, represented by DOC, was swiftly transformed and degraded under the influence of microorganisms, resulting in a substantial decrease in its content (Wang et al. 2016).

3.4.2 Humus content

Humic substances (HS) are types of polymer compounds formed by organic substances under the influence of microorganisms and enzymes, mainly consisting of humic acid (HA) and fulvic acid (FA). In aerobic composting, the goal is to obtain highly humified compost products within a short period. The formation and quantity of HS not only reflect the degree of humification of compost products but also indicate the intensity of biodegradation and the stability of the compost (De Melo et al. 2016). After composting, the HS content in the CK treatment was 154.8 g kg−1, whereas, in the F1, F3, F1 + B15, and F3 + B15 treatments, it significantly decreased to a range of 118.4–138.9 g kg−1 (Fig. 4c).

HA contains more aromatic carbon and represents a macromolecular organic substance soluble in dilute alkali but insoluble in dilute acid (Li et al. 2021). At the end of composting, the content of HA in the F1, F3, F1 + B15, and F3 + B15 treatments decreased by 14.6–31.9% compared to the control, and this difference was statistically significant (P < 0.05). This suggests that the addition of Fe/BC did not favor HA formation. Fe3+ readily chelates with hydroxyl, carboxyl, and other functional groups in HA, leading to the formation of stable chelating structures, which consequently decreases the content of HA in the compost (Karlsson and Persson 2012).

On the other hand, FA, as one of the components of humus substances, possesses a smaller molecular weight, higher activity, and greater degree of oxidation compared to HA (Shi et al. 2020). Contrary to CK, the content of FA in the F1, F3, F1 + B15, and F3 + B15 treatments increased by 2.6–10.9% at the end of composting, but there were no significant differences in FA content among the treatments. This indicates that the addition of Fe/BC had limited effects on the activity of microorganisms involved in FA formation and decomposition.

HA/FA ratio serves as an indicator to assess the degree of maturation. A value greater than 1.7 suggests that the pile is essentially decomposed, with higher values indicating a higher degree of humus in the compost (Shan et al. 2013). In this study, the HA/FA of the F1, F3, F1 + B15, and F3 + B15 treatments decreased by 20.0–38.4% compared to CK, with the lowest HA/FA (2.5) observed in the F3 + B15 treatment (Fig. 4d). These findings indicate that the humification of organic matter was achieved in all treatments, while the addition of Fe/BC had a somewhat negative effect on humification to some extent.

3.4.3 Dissolved organic matter fluorescence spectra analysis

The composting process’s stability can be partially inferred from the changes in DOM content (Smebye et al. 2016). The fluorescence spectrum of DOM in composting was obtained through PARAFAC analysis (Fig. 5a). The results revealed four distinct components: Component 1 (C1) represented terrestrial humus with excitation (Ex) and emission (Em) wavelengths around 365 nm and 464 nm, respectively. Component 2 (C2) displayed wavelengths of approximately 340/402 nm at Ex/Em, resembling marine humus but likely derived from land-based organic matter. Component 3 (C3) shared similarities with aromatic compounds, with a peak Ex/Em value of about 305/400 nm. Component 4 (C4) exhibited similarities with FA (Fulvic Acid) with a peak Ex/Em value of around 420/488 nm. Additionally, quantitative PARAFAC analysis was conducted on composting samples under different treatments (Additional file 1: Fig. S2).

Fig. 5
figure 5

a The fluorescence spectra of dissolved organic matter of composting samples; b the maximum fluorescence intensity of four fluorescent components of dissolved organic matter (Fmax); c the fluorescence parameter of dissolved organic matter. BIX Biotic index, FI fluorescence index, HIX humification index. Data are mean ± SD (n = 3)

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In Fig. 5b, the DOM content of F1, F3, F1 + B15, and F3 + B15 treatments decreased significantly by 29.3–68.8% (P < 0.05) compared to the CK group, with the most substantial decrease observed in the F3 + B15 treatment. This decline may be attributed to Fe/BC promoting the chemical decomposition of DOM or molecular polymerization, leading to the formation of macromolecular humus (López-González et al. 2013).

Furthermore, the addition of FeCl3 and BC influenced the fluorescence parameters of DOM at the end of composting, allowing for a further evaluation of the effect of Fe/BC co-conditioning in composting. The results indicated that BIX and HIX values of DOM in F1, F3, F1 + B15, and F3 + B15 treatments increased significantly by 18.9–25.1% and 4.4–18.5% (P < 0.05) (Fig. 5c) when compared to the CK. Among these treatments, F3 displayed the highest BIX value of 0.65, while F3 + B15 showed the highest HIX value of 1.05. On the other hand, the FI (Fluorescence Index) values in F1, F3, and F1 + B15 treatments showed no significant difference when compared to the CK. The HIX value serves as a measure of the degree of humification of DOM, where higher values indicate a more complex molecular composition with increased contents of humus substances and macromolecular compounds (Halbedel and Herzsprung 2020). This observation reflects that Fe/BC co-conditioner can enhance the degree of DOM humification.

3.5 Nitrification genes abundance during Fe/BC co-conditioning composting

The changes in the expression of nitrification-related genes (AOA and AOB) during the composting process are depicted in Fig. 6. During the high-temperature phase of composting, the copy number of AOA genes was significantly lower in F1, F3, F1 + B15, and F3 + B15 compared to CK (Fig. 6a). This decrease might be attributed to the acidic environment inhibiting the growth and activity of ammonia-oxidizing bacteria. However, by the end of composting, the copy number of AOA genes in F1 + B15 and F3 + B15 was significantly higher than that in CK (P < 0.05). This outcome can be attributed to the porous structure of biochar, which provides a suitable environment for microorganisms to multiply and facilitates NH3 oxidation. Additionally, the abundant acidic functional groups on the biochar surface effectively absorb NH4+ and promote the ammonia oxidation process.

Fig. 6
figure 6

a Copy number of AOA genes and b copy number of AOB genes during the composting process. Data are mean ± SD (n = 3)

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As shown in Fig. 6b, by the end of composting, the copy number of AOB genes in F1, F3, F1 + B15, and F3 + B15 increased by 37.6%, 42.3%, 65.8%, and 170.5%, respectively, compared to the CK treatment. These results indicate that the addition of Fe/BC conditioner enhanced the copy number of nitrifying bacteria, consequently reducing NH3 emissions from the composting process.

3.6 NH 3 reduction mechanisms and correlation analysis

To analyze the relationship between NH3 emissions in Fe/BC co-conditioning composting and properties, a structural equation model analysis was conducted in this study, and the results are presented in Fig. 7a. NH3 emissions were found to be positively correlated with TN (P < 0.001) but negatively correlated with NH4+ (P < 0.001). This can be attributed to the addition of Fe/BC co-conditioner, which absorbs more NH4+ during composting, inhibiting the conversion of NH4+ to NH3. Furthermore, the copy number of AOA genes displayed a significant negative correlation with NH4+ (P < 0.05), while the copy number of AOB genes showed a significant positive correlation with TN (P < 0.001). These results suggest that the increase in TN content in composting, caused by the addition of Fe/BC co-conditioners, is associated with the copy number of AOB genes.

Fig. 7
figure 7

a Structural equation model between NH3 character indexes (the dashed blue line indicates no significance; the solid green line indicates a positive association; the solid yellow line indicates a negative association, * P < 0.05; *** P < 0.001), b schematic diagram of the mechanism of Fe/BC co-conditioning to reduce NH3 emissions

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The potential mechanism underlying the reduction of NH3 emissions through Fe/BC co-conditioning is summarized in Fig. 7b. The hydrolysis of FeCl3 produces a large amount of H+, reducing the pH value of the composting. This, in turn, inhibits microbial activity and decreases the decomposition rate of organic matter, thereby inhibiting NH3 production. Moreover, Fe3+ can react with the reducing groups (-OH, -COOH) on the surface of FA in the composting, forming compounds via a bridging effect that absorbs NH4+ and further reduces NH3 emissions (Nahm 2003).

Additionally, the reduction of NH3 emissions attributed to the addition of biochar is linked to its high specific surface area and porous structure. Agyarko-Mintah et al. (2017) demonstrated that NH3 adsorption mainly occurs in the pore space of biochar. Furthermore, the abundant acid functional groups on the biochar surface effectively adsorb NH4+ and NH3, enhancing nitrification and reducing NH3 emissions (Khan et al. 2020). Moreover, biochar increases cellulase activity during composting, leading to more soluble organic carbon from cellulose decomposition. This, in turn, improves the utilization of NH4+ by microorganisms and ultimately reduces NH3 emissions (Yin et al. 2023).

3.7 Environmental and economic benefit analysis of Fe/BC co-conditioning

Evaluating the benefits of conditioners for composting is important, but there is a lack of relevant research. To compare the advantages and disadvantages of Fe/BC co-conditioners with CK, a preliminary economic benefit analysis was conducted, taking into consideration both costs and income. The inputs involved in the calculation were mainly FeCl3 and the amount of biochar used, while the outputs mainly included enhanced fertilizer value and reduced NH3 emission benefits. The net benefit was determined as the difference between the inputs and outputs.

The costs of FeCl3, biochar, inorganic fertilizer, and reduced NH3 emissions are shown in Additional file 1: Table S3a. Additionally, the economic and environmental benefits of adding Fe/BC co-conditioners compared to CK are shown in Additional file 1: Table S3b. Among the treatments, F1 had the lowest economic benefit at 0.1 $ t−1, while the F3 + B15 treatment had the highest economic benefit at 23.6 $ t−1. This indicates that the economic benefit of using both FeCl3 and biochar is superior to that of using FeCl3 alone. Considering that the total amount of NH3 produced by livestock in China was about 5.5 × 109 kg N (Zhang et al. 2020), the reduction in NH3 achieved by adding Fe/BC co-conditioners ranged from 11.1% to 48.2%. If Fe/BC co-conditioners were widely adopted for treating livestock manure in China, it could theoretically generate an economic and environmental benefit of 0.4–1.2 billion $ year−1.

However, the economic and environmental benefit analysis in this study was relatively preliminary. Firstly, it only considered the effect of Fe/BC conditioners on NH3 reduction, while other greenhouse gases like N2O and CH4 were not taken into account. Secondly, the reduction of NH3 emissions also has potential benefits for human health and ecosystem health, but these economic benefits were not assessed in this study. Furthermore, certain parameters of the composting process, such as the cost of composting raw materials, site fees, and labor costs, were not included due to varying market prices in each country. Future studies should quantify these parameters by collecting relevant data to provide a more comprehensive economic analysis.

4 Future research

This study focused on the relationship between Fe/BC co-conditioners on physicochemical properties and NH3 emissions in aerobic composting. In future studies, further research is needed to investigate the interactions between microbial communities in aerobic composting with the addition of Fe/BC co-conditioners, especially the relationship between Fe/BC co-conditioners on the functional N-converting flora. In addition, the potential environmental risks of Fe/BC co-conditioners should be kept in mind during their application. This study did not discuss the possible increased risk of soil-borne disease transmission by Fe addition and its long-term effects on crop growth, and future analyses of the interactions between soil, microbes and crops with Fe/BC organic manures need to be further strengthened.

5 Conclusion

In this study, we investigated the effects of Fe/BC co-conditioners on NH3 emissions, physical and chemical properties, nutrient dynamics, and humus and organic matter contents. The results of this study show that Fe/BC co-conditioning created an acidic environment that resulted in a significant reduction of cumulative NH3 emissions during the high temperature period of the composting process. Moreover, the addition of Fe/BC co-conditioners facilitated the absorption of NH4+–N and NO3−–N in the composting process through adsorption, effectively inhibiting NH3 emissions. Fe/BC co-conditioning also reacted with the reducing substances in composting, which had a positive effect on the degradation of organic matter and humic acid, and improved the humification index of DOM. Comparative analysis with traditional organic compost demonstrated that the incorporation of Fe/BC co-conditioners generated more substantial economic and environmental benefits. Overall, this study highlights the effectiveness of Fe/BC co-conditioning in reducing NH3 emissions and enhancing composting quality. As a promising approach for the reduction and harmless utilization of manure, it holds tremendous potential for widespread application. However, when Fe/BC organic fertilizers are applied in field trials, there is a need to observe the long-term effects on crops and to carry out risk assessments.