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bentonite on nitrogen transformation and microbial community dynamics during aerobic composting of pig manure
Journal Pre-proofs Beneficial effects of bacterial agent/bentonite on nitrogen transformation and microbial community dynamics during aerobic composting of pig manure Honghong Guo, Jie Gu, Xiaojuan Wang, Mubasher Nasir, Jing Yu, Liusheng Lei, Jia Wang, Wenya Zhao, Xiaoxia Dai PII: DOI: Reference:
S0960-8524(19)31614-1 https://doi.org/10.1016/j.biortech.2019.122384 BITE 122384
To appear in:
Bioresource Technology
Received Date: Revised Date: Accepted Date:
27 September 2019 31 October 2019 1 November 2019
Please cite this article as: Guo, H., Gu, J., Wang, X., Nasir, M., Yu, J., Lei, L., Wang, J., Zhao, W., Dai, X., Beneficial effects of bacterial agent/bentonite on nitrogen transformation and microbial community dynamics during aerobic composting of pig manure, Bioresource Technology (2019), doi: https://doi.org/10.1016/j.biortech.2019.122384
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Title Page
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Title:
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Beneficial effects of bacterial agent/bentonite on nitrogen transformation and
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microbial community dynamics during aerobic composting of pig manure
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Authors:
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Honghong Guoa, Jie Gua,b*, Xiaojuan Wanga, Mubasher Nasira, Jing Yua, Liusheng
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Leia, Jia Wang a, Wenya Zhaoa, Xiaoxia Daia
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Affiliations and addresses:
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a College
of Natural Resources and Environment, Northwest A&F University,
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Yangling, Shaanxi 712100, China
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b
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Province, Northwest A&F University, Yangling, Shaanxi 712100, China
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Corresponding Author: Jie Gu
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Email: [email protected]
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Tel: 86-029-87081265
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Fax: 86-029-87081265
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Address: College of Natural Resources and Environment, Northwest A&F University,
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Yangling, Shaanxi 712100, China
Research Center of Recycle Agricultural Engineering and Technology of Shaanxi
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Abstract:
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This study investigated the effects of adding a bacterial agent (B) and bentonite (BT)
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on nitrogen transformation, nitrogen functional genes, and the microbial community
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dynamics during the aerobic composting of pig manure, as well as their contributions
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to NH3 and N2O emissions. Treatments B, BT, and BT+B reduced the NH3 emissions
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by 31.34%, 18.82%, and 23.67%, respectively, and the N2O emissions by 53.16%,
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72.56%, and 63.41%. N2O and NH3 emissions were strongly related to the functional
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genes. Adding bacterial agent promoted the ammonia oxidation process to reduce
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NH3 emissions, whereas the influence of bentonite on nitrogen conversion was mostly
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related to nirS and nirK in denitrification processes. Nitrification and denitrification
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were dominated by different functional microorganisms in various stages of
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composting, where Proteobacteria comprised the most important denitrifying
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microorganisms. Thus, the additives reduced NH3 and N2O emissions by regulating
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nitrification and denitrification processes, and adding both was highly advantageous.
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Keywords: Bacterial agent, Bacterial community, Bentonite, Composting, Nitrogen
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functional gene
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1. Introduction
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China is one of the largest agricultural countries throughout the world, and thus it
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generates an abundance of biomass resources such as agricultural residues and
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livestock manure. In particular, wheat straw is the main biomass raw material
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obtained from agricultural residues (Jiang et al., 2012), while pig manure is the main
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type of livestock manure, where the total mass produced was about 490 million tons
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in 2015 (NBSC, 2016). Aerobic composting is recognized as an effective method for 2
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treating organic solid waste to produce safe, stable, and nutrient-rich soil amendments
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(Awasthi et al., 2018). However, due to the continuous metabolic activities of
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microorganisms during aerobic composting process, a large amount of nitrogen is lost
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through gas emissions (Kim et al., 2017), which greatly influence the composting
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process and the utilization of the compost products (Wang and Zeng, 2017a), as well
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as causing severe air pollution (Awasthi et al., 2018). Nitrification and denitrification
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by microbes are considered to be the main drivers of the nitrogen cycle (Reinhardt et
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al., 2006; Chen et al., 2017). Ammonia oxidizing bacteria are the major ammonia
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oxidizing microorganisms in the composting process, and they can encode amoA
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genes to facilitate nitrification (Jarvis et al., 2009; Zeng et al., 2011). Maeda et al.
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(2010) reported that N2O emissions are positively correlated with the nosZ gene in the
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denitrifying bacterial community. In addition, nitrogen transformation in various
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environments has been evaluated extensively based on nitrogen functional genes
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(Table 1). Therefore, it is important to study the functional genes related to NH3 and
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N2O emissions during composting.
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Additives are a promising strategy for reducing nitrogen losses and inhibiting
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N2O emissions during composting. Previous studies have shown that additives such as
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zeolite (Wang et al., 2017b), medical stone (Awasthi et al., 2018), and biochar (Mao
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et al., 2018) can effectively reduce NH3 emissions by 13.0–78.5% and N2O emissions
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by 37.0–82.4% during the composting of bioorganic matter or pig manure. It is
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considered that these additives often reduce the nitrogen loss caused by NH3 and N2O
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emissions from compost due to their physical adsorption properties. Bentonite is a
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widely distributed inexpenive mineral, which has the advantages of high ion
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adsorption, porosity, and microbial adhesion, and it is suitable for environmental
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protection (Fernández-Nava et al., 2011). In addition, the composting process 3
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involves complex biotransformation, which is particularly important for nitrogen
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conversion and gas generation during composting (Awasthi et al., 2018; Wang and
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Zeng et al., 2017a). Li et al. (2018) showed that inoculation compost with
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microorganisms can promote the succession of the microbial community. Mao et al.
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(2018) also reported that the addition of bacterial powder to compost improved
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nitrogen storage and the microbial community in the compost. Bacillus megaterium
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has been widely used in agriculture and fertilizer manufacturing to promote the uptake
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of nutrients by crops (El-Sirafy et al., 2006; Rolewicz et al., 2018), which is important
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for the development of green agriculture. However, little is known about the possible
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effects of bentonite and Bacillus megaterium as compost additives, including their
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influence on NH3 and N2O emissions.
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The present study investigated the effects of supplementing compost with
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bentonite and Bacillus megaterium on nitrogen conversion, nitrogen functional genes,
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and the microbial community. In addition, the relationships were determined between
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NH3 and N2O emissions, and related factors such as the bacterial community,
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functional genes, and environmental factors. The results obtained in this study provide
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new insights and theoretical support for the development of safer and more
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environmentally friendly composting techniques.
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2. Materials and Methods
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2.1. Raw materials
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The fresh pig manure and dry wheat straw used in the compost were obtained from
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the livestock and poultry breeding base and crop experiment station at Northwest
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A&F University, China. The ratio of pig manure relative to straw was 6:1 in order to
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obtain a suitable moisture content (60%) and C/N ratio (25) in the compost (Bernal et
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al., 2009). Bentonite was purchased from Weifang Huawei Bentonite Group Co. Ltd, 4
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Shaanxi Province, China. Bacillus megaterium was purchased from China General
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Microbiological Culture Collection (CGMCC) Center, and the CGMCC number was
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1.1870. Before composting, the strain was activated and expanded in culture. The
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cultivated strain was then centrifuged, washed, and resuscitated in sterile distilled
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water. Finally, the concentration of the microbial suspension was adjusted to 108
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CFU/mL for the composting experiment.
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2.2. Experimental setup and sample collection
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A composting experiment was performed for 43 days using a laboratory scale
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composting reactor (effective volume of 130 L). The control (CK) treatment was
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conducted without adding bentonite or bacterial agent. The BT treatment contained
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bentonite added at 5% of the dry weight of the mixture. The B treatment contained
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bacterial agent added at 5% of the dry weight of the mixture. The BT+B treatment
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contained bentonite and bacterial agent added together at 5% of the dry weight of the
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mixture. All treatments were mixed thoroughly and placed in a composting reactor.
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Each composting reactor contained a hole for pumping air via the pump at the bottom
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of the reactor, a hole for measuring the compost temperature in the middle of the
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reactor, and a hole at the top of the reactor to allow gas outflow. A compost
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ventilation rate of 0.35 L/kg (dry weight)/h was maintained during composting (Li et
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al., 2012) and gas samples were collected using gas sampling bags. The compost
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temperature and room temperature were measured three times per day and their
117
averages were determined. In addition, the volatilization of NH3 usually occurs during
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the thermophilic stage of composting, so the N2O emissions are mainly released in the
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initial and/or mature stages of composting. Therefore, we collected solid samples after
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0, 3, 12, 17, and 43 days for further analysis. Before sample collection, the samples
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were turned over to ensure their uniformity. Each collected sample was divided into 5
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two parts, where one was used to determine the physicochemical properties and the
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other for molecular microbial analysis after freeze drying.
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2.3. Compost analysis and gas determination
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The pH of the compost was determined using an aqueous extract of each fresh sample
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(1:10, w / v, sample / Dis. H2O ratio) with a pH electrode (Thermo science, Orion 4
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Star) (Kim et al., 2017). NH4+-N and NO3–-N were extracted with 2 M KCl (1:20) and
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analyzed using a segmented flow analyzer (Technicon Autoanalyzer II system,
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Germany) (Liu et al., 2019). The C/N ratios were determined based on the total
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nitrogen and total carbon contents obtained using a Vario Macro CHNS Element
131
Analyzer. N2O emissions were determined by gas chromatography (Agilent
132
Technologies, China). NH3 was absorbed by H3BO3 and the ammonia content was
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titrated against H2SO4 to determine the NH3 emissions during composting (Yang et al.,
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2015).
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2.4. High-throughput sequencing of microbial communities
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DNA was extracted from 0.1 g of the freeze-dried samples collected on days 0, 3, 12,
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17, and 43 using a Fast DNA Kit for Soil (MP Biomedicals, USA) according to the
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manufacturer’s instructions. The concentration and purity of DNA were measured
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using an Epoch multi-volume spectrophotometer (BioTek, USA).
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Bacterial 16S rRNA was amplified by PCR using the V3–V4 region universal
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primer U341F/U806R and sequenced with the Illumina HiSeq 2500 platform. The
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sequences obtained were submitted to QIIME (http://qiime.org/index.html) for quality
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control and clustered into operational taxonomic units at a shared identity threshold of
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97%. For each representative sequence, the classification information was annotated
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using the RDP classifier-based SILVA database (http://www.arb-silva.de). Functional
6
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bacteria involved in nitrification and denitrification were screened as described
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previously by Ding et al. (2017), Guo et al. (2013), and Shu et al. (2016).
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2.5. Quantitative analysis of nitrogen functional genes
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Real-time quantitative PCR (qPCR) was conducted to quantify five nitrogen
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functional genes comprising bacterial ammonia monooxygenase (amoA), nitrate
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reductase (napA), nitrite reductases (nirK and nirS), and nitrous oxide reductase
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(nosZ), which are involved in nitrogen nitrification and denitrification processes (Guo
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et al., 2019a; Zhang et al., 2018a). The qPCR amplification reaction was conducted as
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follows: (1) initial denaturation for 15 min at 95°C, followed by (2) 40 cycles at 95°C
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for 10 s, annealing for 20 s, and then extension at 72°C for 32 s. The gene copy
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numbers in the samples were calculated using the external standard curve method.
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2.6. Statistical analysis
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Changes in the physicochemical properties, gas emissions, and gene abundances were
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plotted using Origin 8.5. SPSS 20.0 was used to evaluate the correlations and
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significance differences among the treatment. Redundancy analysis (RDA) was
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performed using Canoco 5.0 to determine the potential factors responsible for
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nitrogen transformation.
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3. Results and discussion
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3.1. Changes in physicochemical properties and enhancement of nitrogen storage
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Aerobic composting is a complex biochemical process where the composting
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temperature increases with the mineralization and metabolism of the available organic
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matter by indigenous microorganisms, and thus changes in the temperature reflect the
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overall microbial activity (Bernal et al., 2009). In the present study, the temperature
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bypassed the mesophilic phase in all of the treatments, where it directly reached the
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high temperature stage (> 50C) on day 1, and then passed through the cooling and 7
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maturation stages (Fig. 1a). The peak temperatures in the CK, B, BT, and BT+B
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treatments were 67.0C (day 3), 68.5C (day 3), 68.0C (day 4), and 69.0C (day 3),
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respectively. The temperature changes indicated that the conditions in the composting
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systems were suitable for supporting the activities of microorganisms. Composting in
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the mesophilic phase was maintained for 14 days in the CK and BT+B treatments, and
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15 days in the BT and B treatments. These durations were sufficient to destroy weed
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seeds and pathogenic microorganisms to allow the compost samples to satisfy
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standard hygiene requirements (Bernal et al., 2009; Wang et al., 2016b). As the
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amount of readily available organic matter decreased, the temperature decreased
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steadily in all of the treatments and tended toward ambient levels at the end of
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composting.
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As shown in Fig. 1b, the addition of BT and B increased the initial pH values in
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the compost, where the pH values were 9.28, 9.15, and 9.16 in the B, BT, and BT+B
184
treatments, respectively. Changes in the pH of compost may be related to the
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production of organic acids and the release of ammonia compounds in the composting
186
materials (Wang et al., 2016b). During the first three days of composting, the
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degradation of organic matter results in the production of organic acids, which
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decrease the pH of the compost, before the pH then increases because of the
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mineralization of nitrogen-containing organic matter and the continued decomposition
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of the organic acids. At the end of composting, the pH values in CK, B, BT, and
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BT+B were 9.14, 9.37, 9.21, and 9.11, respectively. The additives had significant
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buffering effects on the changes in the pH during composting, especially in the B
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treatment, which may have been related to the conversion of NH4+-N, and this could
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have further attenuated the impact of the changes in pH on the conversion of NH4+-N
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to NH3, thereby reducing the nitrogen losses (Wang et al., 2017b; Awasthi et al., 8
196 197
2018). Changes in NH4+-N and NO3–-N are important markers of the transformation of
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nitrogen during composting (Fig. 1c and 1d). In all of the treatments, the level of
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NH4+-N did not change significantly in the first three days of composting, before
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decreasing and eventually stabilizing. These changes may be explained by the high
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mineralization and ammonification of organic nitrogen after three days of composting.
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The high temperature and pH in the composting reactor accelerated the conversion of
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NH4+-N into NH3, and these processes were mainly responsible for the loss of
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nitrogen within a short time (Jeong et al., 2017). The NH3 emissions increased rapidly
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in all of the treatments during the early composting stages, before decreasing
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subsequently (Fig. 1e). The NH3 emissions were lower in all of the treatments
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containing additives compared with the CK treatment, where the NH3 emissions from
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B, BT, and BT+B were reduced by 31.34%, 18.82%, and 23.67%, respectively.
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Bentonite has high porosity and a large specific surface area, which can effectively
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adsorb NH3 and NH4+-N, thereby reducing the NH3 emissions during organic solid
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waste composting (Li et al., 2015; Chan et al., 2016). The bacterial agent may have
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reduced the NH3 emissions by promoting the growth of ammonia-oxidizing bacteria,
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which was also supported by the changes in the relative abundance of the amoA gene
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(Fig. 2).
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The conversion of NH4+-N into NH3 decreased and the former tended to
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accumulate during the cooling period of composting. Subsequently, most of the
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NH4+-N was utilized by nitrifying and denitrifying microorganisms (Kim et al., 2017;
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Awasthi et al., 2018), and thus the NO3–-N content was higher in the late composting
219
stage. The great reduction in the NO3–-N contents during the thermophilic period may
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have been due to the inactivation or death of large amounts of nitrifying 9
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microorganisms because of the high temperature (Zhang et al., 2017b). N2O emissions
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were mainly detected in the early and mature stages of composting with high NO3–-N
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contents (Fig. 1f). This may have been due to the rapid degradation of organic matter
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in the early stage, as well as mass sedimentation and water content increases in the
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later stages, which resulted in an insufficient oxygen supply and the formation of
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anaerobic zones in the compost, thereby promoting the conversion of NH4+-N into
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N2O via nitrifying or denitrifying microorganisms (Liu et al., 2019; Awasthi et al.,
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2018). In general, the B, BT, and BT+B treatments reduced the N2O emissions by
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53.16%, 72.56%, and 63.41%, respectively compared with the CK treatment. The
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N2O emissions were mainly attributed to the denitrification process. The compositions
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of the denitrifying bacteria and genes encoding denitrification enzymes (napA, nirK,
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nirS, and nosZ) are closely related to the consumption and production of N2O under
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different environmental conditions (Miao et al., 2015).
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3.2. Changes in the abundances of nitrogen functional genes
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In the present study, denitrifying genes (napA, nirS, nirK, and nosZ) were more
236
abundant than the nitrifying gene (amoA) (Fig. 2), thereby indicating that the
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denitrification process was more active than the nitrification process. A significant
238
increase in all functional genes occurred during the compost cooling period and
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maturity period. A rapid increase occurred in the abundance of napA, followed by
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those of nirK, nirS, and nosZ, which indicates that napA was activated first in the
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denitrification process, before nirK, nirS, and nosZ.
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The bacterial amoA gene is capable of oxidizing ammonia into nitrite, and
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variations in this gene directly influence the NH4+-N and NO3–-N contents during
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composting. The abundance of the amoA gene was lower in the first 17 days and this
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may have been related to the relatively high temperature and low oxygen environment, 10
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which was not favorable for the growth of ammonia-oxidizing bacterial communities
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(Reed et al., 2018). A previous study reported that a temperature range of 15−25°C is
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most suitable for the growth and activity of nitrifying microorganisms (Taylor et al.,
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2017). As the temperature decreased, the nitrifying microbial community recovered
250
and the amoA gene copy numbers peaked at 1.44 106 to 9.66 107 at the end of
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composting. These results are consistent with those obtained previously (Guo et al.,
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2019a; Yan et al., 2018), where the abundance of the bacterial amoA gene decreased
253
to its lowest level in the compost after the continuous thermophilic period, before
254
increasing significantly as the compost entered the maturation stage. The addition of
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the bacterial agent promoted the growth of the amoA gene in the thermophilic phase
256
of composting, thereby indicating that the bacterial agent could promote the oxidation
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of NH4+-N by the amoA gene during the thermophilic phase to reduce NH3 emissions.
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However, the NH4+-N consumption due to ammonia oxidation in the BT treatment
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was much lower than those in the CK and B treatments, and thus the BT and BT+B
260
treatments eventually yielded lower NO3–-N and N2O emissions.
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The napA denitrification gene had the highest abundance in this study. After
262
composting for 3 days, the abundances of napA, nirK, nirS, and nosZ were higher than
263
those in the initial material, and their abundances increased until they peaked in the
264
maturity period. For example, the abundance of the napA gene in the initial material
265
was 9.05 × 106, and it increased to 8.22 109 to 1.34 1010 at the end of composting.
266
These changes may be explained by the composting process promoting the
267
proliferation of denitrifying bacteria (Caceres et al., 2018). The different treatments
268
had significant effects on the denitrification genes. The abundance of the napA gene
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increased rapidly in the B treatment after entering the cooling phase and it eventually
270
exceeded that in the CK treatment. A similar trend was observed for the two 11
271
functionally equivalent nitrite reduction genes (nirS and nirK) in the different
272
treatments during composting. A previous study also showed that the nirS and nirK
273
genes can replace or eliminate each other under certain conditions (Guo et al., 2013).
274
The addition of bentonite promoted increases in the abundances of the nirS and nirK
275
genes, where their abundances exceeded those in the CK treatment at the end of
276
composting. However, the abundance of the nirK gene was higher than that of the
277
nirS gene in the present study. Zhang et al. (2015) also found that the nirK gene is
278
more important than the nirS gene during the composting of agricultural waste.
279
Moreover, the abundance of the nosZ gene was higher in the CK treatment than the
280
other treatments in all of the composting stages, and this difference was more obvious
281
when the compost entered the maturity stage. The production of N2O induces the
282
proliferation of various bacteria that carry the nosZ gene and this might explain the
283
obvious differences among the treatments.
284
3.3. Changes in microbial communities
285
According to the results obtained using the RDP classifier, Firmicutes, Proteobacteria,
286
Actinobacteria, and Bacteroidetes were the dominant phyla in the composting process,
287
where they accounted for 88.23–97.85% of the total sequence reads during
288
composting (Fig. 3a). Firmicutes had the highest abundance in the early stage of
289
composting and throughout the mesophilic period, where this phylum accounted for
290
52.53% of the sequence reads on day 0 and 84.56% on day 3 in the BT treatment, but
291
the abundance gradually decreased during the cooling and maturity stages, where the
292
abundance in the final compost product decreased to 6.02% in BT and 19.59% in CK.
293
As shown in previous studies, Firmicutes was the dominant bacterial phylum in the
294
mid-temperature and thermophilic phases of composting, and members of this phylum
295
have important roles in the utilization and degradation of cellulose (Qian et al., 2016). 12
296
Compared with CK, the additives promoted increases in the abundances of Firmicutes,
297
i.e., by 4.77%, 17.97%, and 14.50% in the B, BT, and BT+B treatments on day 12,
298
and by 48.05%, 89.27%, and 156.63% on day 17. By contrast, the abundance of
299
Proteobacteria was lower in the early and thermophilic periods of composting, but
300
they dominated the cooling and maturity stages. In the compost products,
301
Proteobacteria accounted for 22.93% of the total sequences in BT+B and 34.27% in
302
CK, and their abundances in the B, BT, and BT+B treatments were 30.89%, 16.95%,
303
and 33.08% lower than that in the CK treatment, respectively. Among the
304
nitrogen-converting microorganisms screened in this study, 60.00% belonged to
305
Proteobacteria, thereby indicating that Proteobacteria had profound effects on the
306
migration and transformation of nitrogen in the compost (Chen et al., 2017; Zhang et
307
al., 2018b). Actinobacteria and Bacteroidetes are very sensitive to high temperatures,
308
and the abundances of both were significantly reduced during the thermophilic period,
309
before increasing during the cooling and maturity phases. Compared with CK, the
310
abundances of Actinobacteria in the compost products were 4.56%, 11.74%, and
311
23.06% lower in B, BT, and BT+B, respectively. Bacteroidetes can decompose high
312
molecular weight compounds (Zhu et al., 2017), and the abundance of this phylum
313
increased significantly to become the dominant bacteria during the compost maturity
314
stage, with increases of 181.93%, 190.40%, and 210.42% in B, BT, and BT+B,
315
respectively, compared with the CK treatment.
316
Based on the sequencing results, 40 genera of nitrifying and denitrifying bacteria
317
were screened, which mainly belonged to Proteobacteria, Firmicutes, and
318
Actinobacteria (Fig. 3b). Due to the limitations of the annotation rate for
319
second-generation microbial sequencing (Burke and Darling et al., 2016),
320
Nitrosomonas was the only nitrifying microorganism screened. Previous studies have 13
321
shown that Nitrosomonas are typical ammonia-oxidizing bacteria and they are
322
ubiquitous in systems with high nitrogen removal rates (Siripong and Rittmann, 2007),
323
but they had a very low abundance and only appeared at the end of composting in the
324
present study. By contrast, denitrifying microorganisms were abundant in the compost,
325
where Bacillus, Pseudomonas, and Luteimonas comprised the highly abundant
326
denitrifying microorganisms. The genus Bacillus includes a large number of
327
thermophilic bacteria with high abundances during the thermophilic composting
328
period, and low abundances in the early stage and maturity period (Mitsuhiko et al.,
329
2018). In addition, most of the members of Bacillus have denitrification functions and
330
they can dissimilate and reduce nitrogen compounds (Verbaendert et al., 2011). The
331
bacterial agent increased the abundance of Bacillus during the thermophilic phase in
332
the present study, whereas bentonite decreased its abundance. Compared with the CK
333
treatment, Bacillus increased by 5.9% on day 12 and by 84.4% on day 17 in B, but
334
decreased by 39.53% and 41.53% on day 12 in BT and BT+B, respectively, and by
335
32.44%, and 48.32% on day 17. Pseudomonas includes common denitrifying
336
microorganisms (Julieta et al., 2012), which are mainly present in the thermophilic
337
and maturity stages of composting. Among the other genera detected, Petrimonas can
338
use nitrate as an electron acceptor to promote the denitrification process (Grabowski
339
et al., 2005), and it had a higher abundance during the cooling and final stages of
340
composting, although its abundance was still very low compared with those of other
341
genera. Thus, these results indicate that denitrification was dominated by different
342
denitrifying microorganisms during the various stages of composting.
343
3.4. Factors that affected changes in the bacterial community and nitrogen
344
functional genes
14
345
Most of the environmental factors were indirectly associated with variations in the
346
abundances of genes via their impacts on microorganisms (Guo et al., 2019b).
347
Therefore, the effects of environmental factors on the main bacteria (top 10 phyla)
348
were evaluated by RDA. Overall, the combined parameters in the composting process
349
accounted for 88.5% of the variations in the bacterial communities. Among the
350
variables considered, temperature, N2O, pH, NH4+-N, NH3, and NO3–-N accounted for
351
33.50%, 19.95%, 16.67%, 12.68%, 11.39%, and 5.81% of the total variation,
352
respectively. As shown in Fig. 4(a), the samples obtained from the thermophilic (days
353
3–12), cooling (day 17), and maturity (day 43) periods during composting were
354
significantly separated, which reflected the importance of temperature for the
355
formation of the microbial communities. There were significant differences in the
356
microbial communities under different treatments during the same period, which
357
highlighted the important effects of the additives on the succession of the microbial
358
community. Firmicutes, Tenericutes, Cyanobacteria, and unidentified_Bacteria were
359
the main microorganisms in the thermophilic period. By contrast, Proteobacteria,
360
Actinobacteria, Chloroflexi, Bacteroidetes, Deinococcus-Thermus, and
361
Gemmatimonadetes were the main microorganisms in the cooling and maturity stages.
362
Among the four dominant phyla, there were significant positive correlations between
363
Firmicutes and temperature, NH4+-N, and NH3 (P < 0.05). Proteobacteria,
364
Bacteroidetes, and Actinobacteria had significant positive correlations with N2O (P <
365
0.05). Thus, these dominant species had important roles in nitrogen transformation
366
and release.
367
Furthermore, the correlations were investigated between temperature, pH,
368
nitrogen transformation (NH4+-N, NO3–-N, N2O, and NH3), and nitrogen functional
369
genes (Fig. 4b). Five nitrogen functional genes (amoA, napA, nirS, nirK, and nosZ) 15
370
were positively correlated with pH, NO3–-N, and N2O, but they had negative
371
correlations with temperature, NH4+-N, and NH3. It should be noted that there were
372
significant positive correlations between amoA, nosZ, and NO3–-N (P < 0.01), and
373
significant positive correlations between the five genes and N2O (P < 0.01). In
374
addition, bentonite and the bacterial agent can affect the transformation of nitrogen
375
via different mechanisms. After composting for three days, the CK and B treatments
376
were significantly separated from the BT and BT+B treatments. In the CK and B
377
treatments, nitrogen conversion was mainly related to the amoA-dominated
378
nitrification process, whereas the BT and BT+B treatments were more strongly
379
correlated with denitrification processes dominated by nirS and nirK. In general, the
380
results obtained in the present study suggest that the genes comprising amoA, napA,
381
nirS, nirK, and nosZ had key roles in controlling N2O emissions.
382
4. Conclusion
383
Different additives provide distinct environmental conditions, which can affect
384
bacterial communities and functional genes. In this study, five genes had strong
385
relationships with NH3 and N2O emissions. The enhanced nitrogen transformation
386
with the bacterial agent was related to the nitrification process dominated by the
387
amoA gene. Adding bentonite mainly affected the denitrification process dominated
388
by the nirS and nirK genes. Functional microbial succession occurred during different
389
composting stages, where Proteobacteria comprised the most important denitrifying
390
microorganisms. The additives improved the nitrification and denitrification processes
391
during composting. Applying both additives was more effective at controlling NH3
392
and N2O emissions.
393 394
Note: All of the supplementary data for this study can be found in the online version 16
395
of the article.
396
Acknowledgments
397
This study was funded by the National Natural Science Foundation of China
398
(41671474 and 41601531) and the Science and Technology Plan Key Project of
399
Shaanxi province (2017ZDCXL-SF-03-03).
400
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569
Figure legends:
570
Fig. 1. Changes in temperature, pH, NH4+-N, NO3–-N, NH3, and N2O during
571
composting. CK: no additive; B: 5% added bacterial agent; BT: 5% added
572
bentonite; BT+B: 2.5% added bentonite and 2.5% added bacterial agent.
573
Fig. 2. Changes in the copy numbers of nitrogen conversion-related genes during
574
composting. CK: no additive; B: 5% added bacterial agent; BT: 5% added
575
bentonite; BT+B: 2.5% added bentonite and 2.5% added bacterial agent.
576
Fig. 3. Changes in the bacterial community in response to the addition of bentonite
577
and bacterial agent during composting. (a) Changes in abundance at the bacterial
578
phylum level. (b) Heat map of bacteria associated with nitrogen conversion.
579
Fig. 4. Redundancy analysis of the effects of environmental factors on the changes in
580
the bacterial community and nitrogen conversion-related genes.
581 582 583 584 585 586 587 588 589 590 591 24
592 593 594
Table 1 Abundances of nitrogen functional genes in different environments
Samples resource
Abundance of the genes
Reference
Pig manure aerobic composting
Range from 8.32 105 to 3.36 108 copies/g (bacteria amoA), 3.00 108 to 2.76 109 copies/g (napA), 8.12 107 to 4.17 109 copies/g (nirK), 6.71 107 to 1.08 109 copies/g (nosZ)
Guo et al., 2019a
Cattle manure aerobic composting
Range from 2.25 105 to 2.76 109 copies/g (bacteria amoA)
Yan et al., 2018
Anaerobic digestion of cattle manure
Range from 8.8 108 to 1.2 1010 copies/g (nirS)
Zhang et al., 2018b
Nitrogen-rich grassland soils
Range from 1.6 105 to 1.2 108 copies/g (bacteria amoA)
Di et al., 2009
Temperate steppe soils
Range from 1.49 105 to 1.36 107 copies/g (bacteria amoA), 3.10 109 to 7.97 1010 copies/g (nirS), 7.10 105 to 6.63 106 copies/g (nirK), 1.00 107 to 1.29 108 copies/g (nosZ)
Zhang et al., 2017a
Arable soils
Range from 2.80 × 106 to 3.10 107 copies/g (bacteria amoA), 3.10 107 to 1.90 108 copies/g (archaea amoA)
Wang et al., 2016a
Wastewater treatment systems (wet sludge)
Range from 4.44 106 to 5.79 107 copies/g (bacteria amoA), 3.88 102 to 1.03 106 copies/g (archaea amoA)
Shu et al., 2016
Tannery wastewater treatment plants (wet sludge)
Range from 7.33 103 to 4.58 104 copies per ng DNA (bacteria amoA), 2.72 104 to 2.14 105 copies per ng DNA (nirS), 5.05 103 to 6.02 104 copies per ng DNA (nirK), 2.63 104 to 4.66 105 copies per ng DNA (nosZ),
Wang et al., 2014
595 596 597 598 599 600 601 602 25
603 604 605
Fig. 1.
606 607
Fig. 1. Changes in temperature, pH, NH4+-N, NO3–-N, NH3, and N2O during
608
composting. CK: no additive; B: 5% added bacterial agent; BT: 5% added
609
bentonite; BT+B: 2.5% added bentonite and 2.5% added bacterial agent.
610 611 612 613
26
614 615 616 617
Fig. 2.
618 619
Fig. 2. Changes in the copy numbers of target nitrogen conversion-related genes
620
during composting. CK: no additive; B: 5% added bacterial agent; BT: 5% added
621
bentonite; BT+B: 2.5% added bentonite and 2.5% added bacterial agent.
622 623 624 625 27
626 627 628 629
Fig. 3.
630 631
Fig. 3. Changes in the bacterial community in response to the addition of bentonite
632
and bacterial agent during composting. (a) Changes in abundance at the bacterial
633
phylum level. (b) Heat map of bacteria associated with nitrogen conversion.
634 635 636 28
637 638 639 640
Fig. 4.
641 29
642 643
Fig. 4. Redundancy analysis of the effects of environmental factors on the changes in the bacterial community and nitrogen conversion-related genes.
644
Graphical abstract
645 646
Highlights
647
(1) Bacterial agent and bentonite can effectively reduce NH3 and N2O emissions.
648
(2) Additives enhanced nitrification and denitrification processes in the compost.
649
(3) Bacterial agents promote ammonia oxidation process to reduce NH3 emissions.
650
(4) Bentonite affected denitrification process dominated by nirS and nirK genes.
651
(5) Proteobacteria most important denitrifying microorganisms.
652
30
S0960-8524(19)31614-1 https://doi.org/10.1016/j.biortech.2019.122384 BITE 122384
To appear in:
Bioresource Technology
Received Date: Revised Date: Accepted Date:
27 September 2019 31 October 2019 1 November 2019
Please cite this article as: Guo, H., Gu, J., Wang, X., Nasir, M., Yu, J., Lei, L., Wang, J., Zhao, W., Dai, X., Beneficial effects of bacterial agent/bentonite on nitrogen transformation and microbial community dynamics during aerobic composting of pig manure, Bioresource Technology (2019), doi: https://doi.org/10.1016/j.biortech.2019.122384
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1
Title Page
2
Title:
3
Beneficial effects of bacterial agent/bentonite on nitrogen transformation and
4
microbial community dynamics during aerobic composting of pig manure
5
Authors:
6
Honghong Guoa, Jie Gua,b*, Xiaojuan Wanga, Mubasher Nasira, Jing Yua, Liusheng
7
Leia, Jia Wang a, Wenya Zhaoa, Xiaoxia Daia
8
Affiliations and addresses:
9
a College
of Natural Resources and Environment, Northwest A&F University,
10
Yangling, Shaanxi 712100, China
11
b
12
Province, Northwest A&F University, Yangling, Shaanxi 712100, China
13
Corresponding Author: Jie Gu
14
Email: [email protected]
15
Tel: 86-029-87081265
16
Fax: 86-029-87081265
17
Address: College of Natural Resources and Environment, Northwest A&F University,
18
Yangling, Shaanxi 712100, China
Research Center of Recycle Agricultural Engineering and Technology of Shaanxi
19 20 21 1
22
Abstract:
23
This study investigated the effects of adding a bacterial agent (B) and bentonite (BT)
24
on nitrogen transformation, nitrogen functional genes, and the microbial community
25
dynamics during the aerobic composting of pig manure, as well as their contributions
26
to NH3 and N2O emissions. Treatments B, BT, and BT+B reduced the NH3 emissions
27
by 31.34%, 18.82%, and 23.67%, respectively, and the N2O emissions by 53.16%,
28
72.56%, and 63.41%. N2O and NH3 emissions were strongly related to the functional
29
genes. Adding bacterial agent promoted the ammonia oxidation process to reduce
30
NH3 emissions, whereas the influence of bentonite on nitrogen conversion was mostly
31
related to nirS and nirK in denitrification processes. Nitrification and denitrification
32
were dominated by different functional microorganisms in various stages of
33
composting, where Proteobacteria comprised the most important denitrifying
34
microorganisms. Thus, the additives reduced NH3 and N2O emissions by regulating
35
nitrification and denitrification processes, and adding both was highly advantageous.
36 37
Keywords: Bacterial agent, Bacterial community, Bentonite, Composting, Nitrogen
38
functional gene
39 40
1. Introduction
41
China is one of the largest agricultural countries throughout the world, and thus it
42
generates an abundance of biomass resources such as agricultural residues and
43
livestock manure. In particular, wheat straw is the main biomass raw material
44
obtained from agricultural residues (Jiang et al., 2012), while pig manure is the main
45
type of livestock manure, where the total mass produced was about 490 million tons
46
in 2015 (NBSC, 2016). Aerobic composting is recognized as an effective method for 2
47
treating organic solid waste to produce safe, stable, and nutrient-rich soil amendments
48
(Awasthi et al., 2018). However, due to the continuous metabolic activities of
49
microorganisms during aerobic composting process, a large amount of nitrogen is lost
50
through gas emissions (Kim et al., 2017), which greatly influence the composting
51
process and the utilization of the compost products (Wang and Zeng, 2017a), as well
52
as causing severe air pollution (Awasthi et al., 2018). Nitrification and denitrification
53
by microbes are considered to be the main drivers of the nitrogen cycle (Reinhardt et
54
al., 2006; Chen et al., 2017). Ammonia oxidizing bacteria are the major ammonia
55
oxidizing microorganisms in the composting process, and they can encode amoA
56
genes to facilitate nitrification (Jarvis et al., 2009; Zeng et al., 2011). Maeda et al.
57
(2010) reported that N2O emissions are positively correlated with the nosZ gene in the
58
denitrifying bacterial community. In addition, nitrogen transformation in various
59
environments has been evaluated extensively based on nitrogen functional genes
60
(Table 1). Therefore, it is important to study the functional genes related to NH3 and
61
N2O emissions during composting.
62
Additives are a promising strategy for reducing nitrogen losses and inhibiting
63
N2O emissions during composting. Previous studies have shown that additives such as
64
zeolite (Wang et al., 2017b), medical stone (Awasthi et al., 2018), and biochar (Mao
65
et al., 2018) can effectively reduce NH3 emissions by 13.0–78.5% and N2O emissions
66
by 37.0–82.4% during the composting of bioorganic matter or pig manure. It is
67
considered that these additives often reduce the nitrogen loss caused by NH3 and N2O
68
emissions from compost due to their physical adsorption properties. Bentonite is a
69
widely distributed inexpenive mineral, which has the advantages of high ion
70
adsorption, porosity, and microbial adhesion, and it is suitable for environmental
71
protection (Fernández-Nava et al., 2011). In addition, the composting process 3
72
involves complex biotransformation, which is particularly important for nitrogen
73
conversion and gas generation during composting (Awasthi et al., 2018; Wang and
74
Zeng et al., 2017a). Li et al. (2018) showed that inoculation compost with
75
microorganisms can promote the succession of the microbial community. Mao et al.
76
(2018) also reported that the addition of bacterial powder to compost improved
77
nitrogen storage and the microbial community in the compost. Bacillus megaterium
78
has been widely used in agriculture and fertilizer manufacturing to promote the uptake
79
of nutrients by crops (El-Sirafy et al., 2006; Rolewicz et al., 2018), which is important
80
for the development of green agriculture. However, little is known about the possible
81
effects of bentonite and Bacillus megaterium as compost additives, including their
82
influence on NH3 and N2O emissions.
83
The present study investigated the effects of supplementing compost with
84
bentonite and Bacillus megaterium on nitrogen conversion, nitrogen functional genes,
85
and the microbial community. In addition, the relationships were determined between
86
NH3 and N2O emissions, and related factors such as the bacterial community,
87
functional genes, and environmental factors. The results obtained in this study provide
88
new insights and theoretical support for the development of safer and more
89
environmentally friendly composting techniques.
90
2. Materials and Methods
91
2.1. Raw materials
92
The fresh pig manure and dry wheat straw used in the compost were obtained from
93
the livestock and poultry breeding base and crop experiment station at Northwest
94
A&F University, China. The ratio of pig manure relative to straw was 6:1 in order to
95
obtain a suitable moisture content (60%) and C/N ratio (25) in the compost (Bernal et
96
al., 2009). Bentonite was purchased from Weifang Huawei Bentonite Group Co. Ltd, 4
97
Shaanxi Province, China. Bacillus megaterium was purchased from China General
98
Microbiological Culture Collection (CGMCC) Center, and the CGMCC number was
99
1.1870. Before composting, the strain was activated and expanded in culture. The
100
cultivated strain was then centrifuged, washed, and resuscitated in sterile distilled
101
water. Finally, the concentration of the microbial suspension was adjusted to 108
102
CFU/mL for the composting experiment.
103
2.2. Experimental setup and sample collection
104
A composting experiment was performed for 43 days using a laboratory scale
105
composting reactor (effective volume of 130 L). The control (CK) treatment was
106
conducted without adding bentonite or bacterial agent. The BT treatment contained
107
bentonite added at 5% of the dry weight of the mixture. The B treatment contained
108
bacterial agent added at 5% of the dry weight of the mixture. The BT+B treatment
109
contained bentonite and bacterial agent added together at 5% of the dry weight of the
110
mixture. All treatments were mixed thoroughly and placed in a composting reactor.
111
Each composting reactor contained a hole for pumping air via the pump at the bottom
112
of the reactor, a hole for measuring the compost temperature in the middle of the
113
reactor, and a hole at the top of the reactor to allow gas outflow. A compost
114
ventilation rate of 0.35 L/kg (dry weight)/h was maintained during composting (Li et
115
al., 2012) and gas samples were collected using gas sampling bags. The compost
116
temperature and room temperature were measured three times per day and their
117
averages were determined. In addition, the volatilization of NH3 usually occurs during
118
the thermophilic stage of composting, so the N2O emissions are mainly released in the
119
initial and/or mature stages of composting. Therefore, we collected solid samples after
120
0, 3, 12, 17, and 43 days for further analysis. Before sample collection, the samples
121
were turned over to ensure their uniformity. Each collected sample was divided into 5
122
two parts, where one was used to determine the physicochemical properties and the
123
other for molecular microbial analysis after freeze drying.
124
2.3. Compost analysis and gas determination
125
The pH of the compost was determined using an aqueous extract of each fresh sample
126
(1:10, w / v, sample / Dis. H2O ratio) with a pH electrode (Thermo science, Orion 4
127
Star) (Kim et al., 2017). NH4+-N and NO3–-N were extracted with 2 M KCl (1:20) and
128
analyzed using a segmented flow analyzer (Technicon Autoanalyzer II system,
129
Germany) (Liu et al., 2019). The C/N ratios were determined based on the total
130
nitrogen and total carbon contents obtained using a Vario Macro CHNS Element
131
Analyzer. N2O emissions were determined by gas chromatography (Agilent
132
Technologies, China). NH3 was absorbed by H3BO3 and the ammonia content was
133
titrated against H2SO4 to determine the NH3 emissions during composting (Yang et al.,
134
2015).
135
2.4. High-throughput sequencing of microbial communities
136
DNA was extracted from 0.1 g of the freeze-dried samples collected on days 0, 3, 12,
137
17, and 43 using a Fast DNA Kit for Soil (MP Biomedicals, USA) according to the
138
manufacturer’s instructions. The concentration and purity of DNA were measured
139
using an Epoch multi-volume spectrophotometer (BioTek, USA).
140
Bacterial 16S rRNA was amplified by PCR using the V3–V4 region universal
141
primer U341F/U806R and sequenced with the Illumina HiSeq 2500 platform. The
142
sequences obtained were submitted to QIIME (http://qiime.org/index.html) for quality
143
control and clustered into operational taxonomic units at a shared identity threshold of
144
97%. For each representative sequence, the classification information was annotated
145
using the RDP classifier-based SILVA database (http://www.arb-silva.de). Functional
6
146
bacteria involved in nitrification and denitrification were screened as described
147
previously by Ding et al. (2017), Guo et al. (2013), and Shu et al. (2016).
148
2.5. Quantitative analysis of nitrogen functional genes
149
Real-time quantitative PCR (qPCR) was conducted to quantify five nitrogen
150
functional genes comprising bacterial ammonia monooxygenase (amoA), nitrate
151
reductase (napA), nitrite reductases (nirK and nirS), and nitrous oxide reductase
152
(nosZ), which are involved in nitrogen nitrification and denitrification processes (Guo
153
et al., 2019a; Zhang et al., 2018a). The qPCR amplification reaction was conducted as
154
follows: (1) initial denaturation for 15 min at 95°C, followed by (2) 40 cycles at 95°C
155
for 10 s, annealing for 20 s, and then extension at 72°C for 32 s. The gene copy
156
numbers in the samples were calculated using the external standard curve method.
157
2.6. Statistical analysis
158
Changes in the physicochemical properties, gas emissions, and gene abundances were
159
plotted using Origin 8.5. SPSS 20.0 was used to evaluate the correlations and
160
significance differences among the treatment. Redundancy analysis (RDA) was
161
performed using Canoco 5.0 to determine the potential factors responsible for
162
nitrogen transformation.
163
3. Results and discussion
164
3.1. Changes in physicochemical properties and enhancement of nitrogen storage
165
Aerobic composting is a complex biochemical process where the composting
166
temperature increases with the mineralization and metabolism of the available organic
167
matter by indigenous microorganisms, and thus changes in the temperature reflect the
168
overall microbial activity (Bernal et al., 2009). In the present study, the temperature
169
bypassed the mesophilic phase in all of the treatments, where it directly reached the
170
high temperature stage (> 50C) on day 1, and then passed through the cooling and 7
171
maturation stages (Fig. 1a). The peak temperatures in the CK, B, BT, and BT+B
172
treatments were 67.0C (day 3), 68.5C (day 3), 68.0C (day 4), and 69.0C (day 3),
173
respectively. The temperature changes indicated that the conditions in the composting
174
systems were suitable for supporting the activities of microorganisms. Composting in
175
the mesophilic phase was maintained for 14 days in the CK and BT+B treatments, and
176
15 days in the BT and B treatments. These durations were sufficient to destroy weed
177
seeds and pathogenic microorganisms to allow the compost samples to satisfy
178
standard hygiene requirements (Bernal et al., 2009; Wang et al., 2016b). As the
179
amount of readily available organic matter decreased, the temperature decreased
180
steadily in all of the treatments and tended toward ambient levels at the end of
181
composting.
182
As shown in Fig. 1b, the addition of BT and B increased the initial pH values in
183
the compost, where the pH values were 9.28, 9.15, and 9.16 in the B, BT, and BT+B
184
treatments, respectively. Changes in the pH of compost may be related to the
185
production of organic acids and the release of ammonia compounds in the composting
186
materials (Wang et al., 2016b). During the first three days of composting, the
187
degradation of organic matter results in the production of organic acids, which
188
decrease the pH of the compost, before the pH then increases because of the
189
mineralization of nitrogen-containing organic matter and the continued decomposition
190
of the organic acids. At the end of composting, the pH values in CK, B, BT, and
191
BT+B were 9.14, 9.37, 9.21, and 9.11, respectively. The additives had significant
192
buffering effects on the changes in the pH during composting, especially in the B
193
treatment, which may have been related to the conversion of NH4+-N, and this could
194
have further attenuated the impact of the changes in pH on the conversion of NH4+-N
195
to NH3, thereby reducing the nitrogen losses (Wang et al., 2017b; Awasthi et al., 8
196 197
2018). Changes in NH4+-N and NO3–-N are important markers of the transformation of
198
nitrogen during composting (Fig. 1c and 1d). In all of the treatments, the level of
199
NH4+-N did not change significantly in the first three days of composting, before
200
decreasing and eventually stabilizing. These changes may be explained by the high
201
mineralization and ammonification of organic nitrogen after three days of composting.
202
The high temperature and pH in the composting reactor accelerated the conversion of
203
NH4+-N into NH3, and these processes were mainly responsible for the loss of
204
nitrogen within a short time (Jeong et al., 2017). The NH3 emissions increased rapidly
205
in all of the treatments during the early composting stages, before decreasing
206
subsequently (Fig. 1e). The NH3 emissions were lower in all of the treatments
207
containing additives compared with the CK treatment, where the NH3 emissions from
208
B, BT, and BT+B were reduced by 31.34%, 18.82%, and 23.67%, respectively.
209
Bentonite has high porosity and a large specific surface area, which can effectively
210
adsorb NH3 and NH4+-N, thereby reducing the NH3 emissions during organic solid
211
waste composting (Li et al., 2015; Chan et al., 2016). The bacterial agent may have
212
reduced the NH3 emissions by promoting the growth of ammonia-oxidizing bacteria,
213
which was also supported by the changes in the relative abundance of the amoA gene
214
(Fig. 2).
215
The conversion of NH4+-N into NH3 decreased and the former tended to
216
accumulate during the cooling period of composting. Subsequently, most of the
217
NH4+-N was utilized by nitrifying and denitrifying microorganisms (Kim et al., 2017;
218
Awasthi et al., 2018), and thus the NO3–-N content was higher in the late composting
219
stage. The great reduction in the NO3–-N contents during the thermophilic period may
220
have been due to the inactivation or death of large amounts of nitrifying 9
221
microorganisms because of the high temperature (Zhang et al., 2017b). N2O emissions
222
were mainly detected in the early and mature stages of composting with high NO3–-N
223
contents (Fig. 1f). This may have been due to the rapid degradation of organic matter
224
in the early stage, as well as mass sedimentation and water content increases in the
225
later stages, which resulted in an insufficient oxygen supply and the formation of
226
anaerobic zones in the compost, thereby promoting the conversion of NH4+-N into
227
N2O via nitrifying or denitrifying microorganisms (Liu et al., 2019; Awasthi et al.,
228
2018). In general, the B, BT, and BT+B treatments reduced the N2O emissions by
229
53.16%, 72.56%, and 63.41%, respectively compared with the CK treatment. The
230
N2O emissions were mainly attributed to the denitrification process. The compositions
231
of the denitrifying bacteria and genes encoding denitrification enzymes (napA, nirK,
232
nirS, and nosZ) are closely related to the consumption and production of N2O under
233
different environmental conditions (Miao et al., 2015).
234
3.2. Changes in the abundances of nitrogen functional genes
235
In the present study, denitrifying genes (napA, nirS, nirK, and nosZ) were more
236
abundant than the nitrifying gene (amoA) (Fig. 2), thereby indicating that the
237
denitrification process was more active than the nitrification process. A significant
238
increase in all functional genes occurred during the compost cooling period and
239
maturity period. A rapid increase occurred in the abundance of napA, followed by
240
those of nirK, nirS, and nosZ, which indicates that napA was activated first in the
241
denitrification process, before nirK, nirS, and nosZ.
242
The bacterial amoA gene is capable of oxidizing ammonia into nitrite, and
243
variations in this gene directly influence the NH4+-N and NO3–-N contents during
244
composting. The abundance of the amoA gene was lower in the first 17 days and this
245
may have been related to the relatively high temperature and low oxygen environment, 10
246
which was not favorable for the growth of ammonia-oxidizing bacterial communities
247
(Reed et al., 2018). A previous study reported that a temperature range of 15−25°C is
248
most suitable for the growth and activity of nitrifying microorganisms (Taylor et al.,
249
2017). As the temperature decreased, the nitrifying microbial community recovered
250
and the amoA gene copy numbers peaked at 1.44 106 to 9.66 107 at the end of
251
composting. These results are consistent with those obtained previously (Guo et al.,
252
2019a; Yan et al., 2018), where the abundance of the bacterial amoA gene decreased
253
to its lowest level in the compost after the continuous thermophilic period, before
254
increasing significantly as the compost entered the maturation stage. The addition of
255
the bacterial agent promoted the growth of the amoA gene in the thermophilic phase
256
of composting, thereby indicating that the bacterial agent could promote the oxidation
257
of NH4+-N by the amoA gene during the thermophilic phase to reduce NH3 emissions.
258
However, the NH4+-N consumption due to ammonia oxidation in the BT treatment
259
was much lower than those in the CK and B treatments, and thus the BT and BT+B
260
treatments eventually yielded lower NO3–-N and N2O emissions.
261
The napA denitrification gene had the highest abundance in this study. After
262
composting for 3 days, the abundances of napA, nirK, nirS, and nosZ were higher than
263
those in the initial material, and their abundances increased until they peaked in the
264
maturity period. For example, the abundance of the napA gene in the initial material
265
was 9.05 × 106, and it increased to 8.22 109 to 1.34 1010 at the end of composting.
266
These changes may be explained by the composting process promoting the
267
proliferation of denitrifying bacteria (Caceres et al., 2018). The different treatments
268
had significant effects on the denitrification genes. The abundance of the napA gene
269
increased rapidly in the B treatment after entering the cooling phase and it eventually
270
exceeded that in the CK treatment. A similar trend was observed for the two 11
271
functionally equivalent nitrite reduction genes (nirS and nirK) in the different
272
treatments during composting. A previous study also showed that the nirS and nirK
273
genes can replace or eliminate each other under certain conditions (Guo et al., 2013).
274
The addition of bentonite promoted increases in the abundances of the nirS and nirK
275
genes, where their abundances exceeded those in the CK treatment at the end of
276
composting. However, the abundance of the nirK gene was higher than that of the
277
nirS gene in the present study. Zhang et al. (2015) also found that the nirK gene is
278
more important than the nirS gene during the composting of agricultural waste.
279
Moreover, the abundance of the nosZ gene was higher in the CK treatment than the
280
other treatments in all of the composting stages, and this difference was more obvious
281
when the compost entered the maturity stage. The production of N2O induces the
282
proliferation of various bacteria that carry the nosZ gene and this might explain the
283
obvious differences among the treatments.
284
3.3. Changes in microbial communities
285
According to the results obtained using the RDP classifier, Firmicutes, Proteobacteria,
286
Actinobacteria, and Bacteroidetes were the dominant phyla in the composting process,
287
where they accounted for 88.23–97.85% of the total sequence reads during
288
composting (Fig. 3a). Firmicutes had the highest abundance in the early stage of
289
composting and throughout the mesophilic period, where this phylum accounted for
290
52.53% of the sequence reads on day 0 and 84.56% on day 3 in the BT treatment, but
291
the abundance gradually decreased during the cooling and maturity stages, where the
292
abundance in the final compost product decreased to 6.02% in BT and 19.59% in CK.
293
As shown in previous studies, Firmicutes was the dominant bacterial phylum in the
294
mid-temperature and thermophilic phases of composting, and members of this phylum
295
have important roles in the utilization and degradation of cellulose (Qian et al., 2016). 12
296
Compared with CK, the additives promoted increases in the abundances of Firmicutes,
297
i.e., by 4.77%, 17.97%, and 14.50% in the B, BT, and BT+B treatments on day 12,
298
and by 48.05%, 89.27%, and 156.63% on day 17. By contrast, the abundance of
299
Proteobacteria was lower in the early and thermophilic periods of composting, but
300
they dominated the cooling and maturity stages. In the compost products,
301
Proteobacteria accounted for 22.93% of the total sequences in BT+B and 34.27% in
302
CK, and their abundances in the B, BT, and BT+B treatments were 30.89%, 16.95%,
303
and 33.08% lower than that in the CK treatment, respectively. Among the
304
nitrogen-converting microorganisms screened in this study, 60.00% belonged to
305
Proteobacteria, thereby indicating that Proteobacteria had profound effects on the
306
migration and transformation of nitrogen in the compost (Chen et al., 2017; Zhang et
307
al., 2018b). Actinobacteria and Bacteroidetes are very sensitive to high temperatures,
308
and the abundances of both were significantly reduced during the thermophilic period,
309
before increasing during the cooling and maturity phases. Compared with CK, the
310
abundances of Actinobacteria in the compost products were 4.56%, 11.74%, and
311
23.06% lower in B, BT, and BT+B, respectively. Bacteroidetes can decompose high
312
molecular weight compounds (Zhu et al., 2017), and the abundance of this phylum
313
increased significantly to become the dominant bacteria during the compost maturity
314
stage, with increases of 181.93%, 190.40%, and 210.42% in B, BT, and BT+B,
315
respectively, compared with the CK treatment.
316
Based on the sequencing results, 40 genera of nitrifying and denitrifying bacteria
317
were screened, which mainly belonged to Proteobacteria, Firmicutes, and
318
Actinobacteria (Fig. 3b). Due to the limitations of the annotation rate for
319
second-generation microbial sequencing (Burke and Darling et al., 2016),
320
Nitrosomonas was the only nitrifying microorganism screened. Previous studies have 13
321
shown that Nitrosomonas are typical ammonia-oxidizing bacteria and they are
322
ubiquitous in systems with high nitrogen removal rates (Siripong and Rittmann, 2007),
323
but they had a very low abundance and only appeared at the end of composting in the
324
present study. By contrast, denitrifying microorganisms were abundant in the compost,
325
where Bacillus, Pseudomonas, and Luteimonas comprised the highly abundant
326
denitrifying microorganisms. The genus Bacillus includes a large number of
327
thermophilic bacteria with high abundances during the thermophilic composting
328
period, and low abundances in the early stage and maturity period (Mitsuhiko et al.,
329
2018). In addition, most of the members of Bacillus have denitrification functions and
330
they can dissimilate and reduce nitrogen compounds (Verbaendert et al., 2011). The
331
bacterial agent increased the abundance of Bacillus during the thermophilic phase in
332
the present study, whereas bentonite decreased its abundance. Compared with the CK
333
treatment, Bacillus increased by 5.9% on day 12 and by 84.4% on day 17 in B, but
334
decreased by 39.53% and 41.53% on day 12 in BT and BT+B, respectively, and by
335
32.44%, and 48.32% on day 17. Pseudomonas includes common denitrifying
336
microorganisms (Julieta et al., 2012), which are mainly present in the thermophilic
337
and maturity stages of composting. Among the other genera detected, Petrimonas can
338
use nitrate as an electron acceptor to promote the denitrification process (Grabowski
339
et al., 2005), and it had a higher abundance during the cooling and final stages of
340
composting, although its abundance was still very low compared with those of other
341
genera. Thus, these results indicate that denitrification was dominated by different
342
denitrifying microorganisms during the various stages of composting.
343
3.4. Factors that affected changes in the bacterial community and nitrogen
344
functional genes
14
345
Most of the environmental factors were indirectly associated with variations in the
346
abundances of genes via their impacts on microorganisms (Guo et al., 2019b).
347
Therefore, the effects of environmental factors on the main bacteria (top 10 phyla)
348
were evaluated by RDA. Overall, the combined parameters in the composting process
349
accounted for 88.5% of the variations in the bacterial communities. Among the
350
variables considered, temperature, N2O, pH, NH4+-N, NH3, and NO3–-N accounted for
351
33.50%, 19.95%, 16.67%, 12.68%, 11.39%, and 5.81% of the total variation,
352
respectively. As shown in Fig. 4(a), the samples obtained from the thermophilic (days
353
3–12), cooling (day 17), and maturity (day 43) periods during composting were
354
significantly separated, which reflected the importance of temperature for the
355
formation of the microbial communities. There were significant differences in the
356
microbial communities under different treatments during the same period, which
357
highlighted the important effects of the additives on the succession of the microbial
358
community. Firmicutes, Tenericutes, Cyanobacteria, and unidentified_Bacteria were
359
the main microorganisms in the thermophilic period. By contrast, Proteobacteria,
360
Actinobacteria, Chloroflexi, Bacteroidetes, Deinococcus-Thermus, and
361
Gemmatimonadetes were the main microorganisms in the cooling and maturity stages.
362
Among the four dominant phyla, there were significant positive correlations between
363
Firmicutes and temperature, NH4+-N, and NH3 (P < 0.05). Proteobacteria,
364
Bacteroidetes, and Actinobacteria had significant positive correlations with N2O (P <
365
0.05). Thus, these dominant species had important roles in nitrogen transformation
366
and release.
367
Furthermore, the correlations were investigated between temperature, pH,
368
nitrogen transformation (NH4+-N, NO3–-N, N2O, and NH3), and nitrogen functional
369
genes (Fig. 4b). Five nitrogen functional genes (amoA, napA, nirS, nirK, and nosZ) 15
370
were positively correlated with pH, NO3–-N, and N2O, but they had negative
371
correlations with temperature, NH4+-N, and NH3. It should be noted that there were
372
significant positive correlations between amoA, nosZ, and NO3–-N (P < 0.01), and
373
significant positive correlations between the five genes and N2O (P < 0.01). In
374
addition, bentonite and the bacterial agent can affect the transformation of nitrogen
375
via different mechanisms. After composting for three days, the CK and B treatments
376
were significantly separated from the BT and BT+B treatments. In the CK and B
377
treatments, nitrogen conversion was mainly related to the amoA-dominated
378
nitrification process, whereas the BT and BT+B treatments were more strongly
379
correlated with denitrification processes dominated by nirS and nirK. In general, the
380
results obtained in the present study suggest that the genes comprising amoA, napA,
381
nirS, nirK, and nosZ had key roles in controlling N2O emissions.
382
4. Conclusion
383
Different additives provide distinct environmental conditions, which can affect
384
bacterial communities and functional genes. In this study, five genes had strong
385
relationships with NH3 and N2O emissions. The enhanced nitrogen transformation
386
with the bacterial agent was related to the nitrification process dominated by the
387
amoA gene. Adding bentonite mainly affected the denitrification process dominated
388
by the nirS and nirK genes. Functional microbial succession occurred during different
389
composting stages, where Proteobacteria comprised the most important denitrifying
390
microorganisms. The additives improved the nitrification and denitrification processes
391
during composting. Applying both additives was more effective at controlling NH3
392
and N2O emissions.
393 394
Note: All of the supplementary data for this study can be found in the online version 16
395
of the article.
396
Acknowledgments
397
This study was funded by the National Natural Science Foundation of China
398
(41671474 and 41601531) and the Science and Technology Plan Key Project of
399
Shaanxi province (2017ZDCXL-SF-03-03).
400
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401
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Figure legends:
570
Fig. 1. Changes in temperature, pH, NH4+-N, NO3–-N, NH3, and N2O during
571
composting. CK: no additive; B: 5% added bacterial agent; BT: 5% added
572
bentonite; BT+B: 2.5% added bentonite and 2.5% added bacterial agent.
573
Fig. 2. Changes in the copy numbers of nitrogen conversion-related genes during
574
composting. CK: no additive; B: 5% added bacterial agent; BT: 5% added
575
bentonite; BT+B: 2.5% added bentonite and 2.5% added bacterial agent.
576
Fig. 3. Changes in the bacterial community in response to the addition of bentonite
577
and bacterial agent during composting. (a) Changes in abundance at the bacterial
578
phylum level. (b) Heat map of bacteria associated with nitrogen conversion.
579
Fig. 4. Redundancy analysis of the effects of environmental factors on the changes in
580
the bacterial community and nitrogen conversion-related genes.
581 582 583 584 585 586 587 588 589 590 591 24
592 593 594
Table 1 Abundances of nitrogen functional genes in different environments
Samples resource
Abundance of the genes
Reference
Pig manure aerobic composting
Range from 8.32 105 to 3.36 108 copies/g (bacteria amoA), 3.00 108 to 2.76 109 copies/g (napA), 8.12 107 to 4.17 109 copies/g (nirK), 6.71 107 to 1.08 109 copies/g (nosZ)
Guo et al., 2019a
Cattle manure aerobic composting
Range from 2.25 105 to 2.76 109 copies/g (bacteria amoA)
Yan et al., 2018
Anaerobic digestion of cattle manure
Range from 8.8 108 to 1.2 1010 copies/g (nirS)
Zhang et al., 2018b
Nitrogen-rich grassland soils
Range from 1.6 105 to 1.2 108 copies/g (bacteria amoA)
Di et al., 2009
Temperate steppe soils
Range from 1.49 105 to 1.36 107 copies/g (bacteria amoA), 3.10 109 to 7.97 1010 copies/g (nirS), 7.10 105 to 6.63 106 copies/g (nirK), 1.00 107 to 1.29 108 copies/g (nosZ)
Zhang et al., 2017a
Arable soils
Range from 2.80 × 106 to 3.10 107 copies/g (bacteria amoA), 3.10 107 to 1.90 108 copies/g (archaea amoA)
Wang et al., 2016a
Wastewater treatment systems (wet sludge)
Range from 4.44 106 to 5.79 107 copies/g (bacteria amoA), 3.88 102 to 1.03 106 copies/g (archaea amoA)
Shu et al., 2016
Tannery wastewater treatment plants (wet sludge)
Range from 7.33 103 to 4.58 104 copies per ng DNA (bacteria amoA), 2.72 104 to 2.14 105 copies per ng DNA (nirS), 5.05 103 to 6.02 104 copies per ng DNA (nirK), 2.63 104 to 4.66 105 copies per ng DNA (nosZ),
Wang et al., 2014
595 596 597 598 599 600 601 602 25
603 604 605
Fig. 1.
606 607
Fig. 1. Changes in temperature, pH, NH4+-N, NO3–-N, NH3, and N2O during
608
composting. CK: no additive; B: 5% added bacterial agent; BT: 5% added
609
bentonite; BT+B: 2.5% added bentonite and 2.5% added bacterial agent.
610 611 612 613
26
614 615 616 617
Fig. 2.
618 619
Fig. 2. Changes in the copy numbers of target nitrogen conversion-related genes
620
during composting. CK: no additive; B: 5% added bacterial agent; BT: 5% added
621
bentonite; BT+B: 2.5% added bentonite and 2.5% added bacterial agent.
622 623 624 625 27
626 627 628 629
Fig. 3.
630 631
Fig. 3. Changes in the bacterial community in response to the addition of bentonite
632
and bacterial agent during composting. (a) Changes in abundance at the bacterial
633
phylum level. (b) Heat map of bacteria associated with nitrogen conversion.
634 635 636 28
637 638 639 640
Fig. 4.
641 29
642 643
Fig. 4. Redundancy analysis of the effects of environmental factors on the changes in the bacterial community and nitrogen conversion-related genes.
644
Graphical abstract
645 646
Highlights
647
(1) Bacterial agent and bentonite can effectively reduce NH3 and N2O emissions.
648
(2) Additives enhanced nitrification and denitrification processes in the compost.
649
(3) Bacterial agents promote ammonia oxidation process to reduce NH3 emissions.
650
(4) Bentonite affected denitrification process dominated by nirS and nirK genes.
651
(5) Proteobacteria most important denitrifying microorganisms.
652
30