Influence of medical stone amendment on gaseous emissions, microbial biomass and abundance of ammonia oxidizing bacteria genes during biosolids composting

Bioresource Technology 247 (2018) 970–979

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Influence of medical stone amendment on gaseous emissions, microbial biomass and abundance of ammonia oxidizing bacteria genes during biosolids composting

MARK

Mukesh Kumar Awasthia,b, Quan Wanga, Sanjeev Kumar Awasthia, Meijing Wanga, ⁎ Hongyu Chena, Xiuna Rena, Junchao Zhaoa, Zengqiang Zhanga, a b

College of Natural Resources and Environment, Northwest A & F University, Yangling, Shaanxi Province 712100, PR China Department of Biotechnology, Amicable Knowledge Solution University, Satna, India

A R T I C L E I N F O

A B S T R A C T

Keywords: Biosolids Mineralization Abundance Bacteria genes Microbial biomass

This study aimed to evaluate the feasibility of medical stone (MS) on microbial biomass, bacteria genes copy numbers, mitigation of gaseous emissions and its correlation with analyzed parameters during the biosolids composting. Composting of the biosolids by amendment of MS 0%, 2%, 4%, 6% and 10% (on dry weight basis) was performed using a 130-L composting reactor. The results showed that with increasing the dosage of MS, the CH4, N2O and NH3 emission were reduced by 60.5–88.3%, 46.6–82.4% and 38.2–78.5%, respectively. In addition, the 6–10% MS amendment enhanced the organic waste mineralization and prolonged the thermophilic phase. The abundance of ammonia oxidizing bacteria (AOB) and archaea (AOAB) were decreased during the first 28 days, but considerable increment was observed during the maturation phase which indicated that AOB and AOAB were liable for nitrification during the curing phase of composting. A significant correlation was observed among the all analyzed parameters in 6–10% MS blended treatments.

1. Introduction Huge quantity of biosolids or sewage sludge (SS) is generated from the waste water treatment plant and its improper disposal may cause serious health hazardous and ecological problem (Cai et al., 2016). Many ecofriendly strategies such as anaerobic digestion and composting, since long time are available to managed this kind of organic waste, which is fit with 3R’s (Reduce, Reuse and Recycle) philosophy, and produced stabilized end product (Villasenor et al., 2011; MauliniDuran et al., 2013; Awasthi et al., 2016a, 2017a). However, the recycling of high moisture contain biosolids through anaerobic digestion is one of best ecofriendly disposal methods but the digested residues are most of the time not completely stabilized. And anaerobic digestion has lead several advantages like energy conservation than other technologies but required high level of investment and proper monitoring for commercial scale, and if the process not runs efficiently that can generates less quantity of gases and then noxious odors problems in surrounding area (Teglia et al., 2011; Zeng et al., 2012; Jiang et al., 2016). In addition, it is difficult to store and transport the un-digested residue owing to its high moisture content and odors as well as limit the direct land applications. Therefore, recycling of this un-digested residue

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requires further treatment and has become urgent need for the management of organic residues of biosolids. Composting is cost effective green technology for disposal of biosolids into available stable humus by the influence of microorganism. Although, this technology is commonly applied over the last many years by rapid increases in biosolids production and abundant agricultural resources (Villasenor et al., 2011; Zeng et al., 2014; Malinska et al., 2014). An infelicitous accident during conventional biosolids composting is the depletion of extensive quantity of organic nutrients through gaseous emission (Bong et al., 2016; Awasthi et al., 2017b) and high concentration of heavy metals, which is not only affect the agronomic value of the end product but also involves for global warming and formation of odor as well as phytotoxicity (Fang and Wong, 1999; He et al., 2000; El Kader et al., 2007; Czekala et al., 2016). However, carbon dioxide (CO2) and ammonia (NH3) emission are two important biogenic gases produced due to the rapid mineralization of organic matters but high moisture content reduced the total aerobic microbial abundance and thus decreased the rate of organic matter (OM) mineralization, and produced bulk of CH4 and N2O gases. Simultaneously, out of total gaseous generation during the composting, only 24% CH4 and 3% N2O alone while major amount of (63%) is CO2 emitted,

Corresponding author. E-mail address: [email protected] (Z. Zhang).

http://dx.doi.org/10.1016/j.biortech.2017.09.201 Received 22 August 2017; Received in revised form 26 September 2017; Accepted 28 September 2017 Available online 04 October 2017 0960-8524/ © 2017 Elsevier Ltd. All rights reserved.

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properly in a separate container on day 0, 3, 7, 10, 14, 21, 28, 35 and 42; while periodically moisture content of the composting mixture was readjusted to ∼55% on turning days. On each mixing days (0, 3, 7, 10, 14, 21, 28, 35 and 42), ∼250.0 g compost samples were taken, which were split into two parts and preserved at 4 and −20 °C, respectively. The composting mass temperature was observed every day four times (6 h) to evaluate the progressiveness of the process and ventilation according to our previous study (Awasthi et al., 2016a).

respectively (Beck-Friis et al., 2000; Fukumoto et al., 2003; SanchezMonedero et al., 2010; Wang et al., 2016); but CH4 and N2O global warming effect 30 and 210 times higher than CO2 (Sommer and Moller 2000; Luo et al., 2013; Sanchez-Garcia et al., 2015). Consequently, SS normally has low C/N ratio that were negatively influenced the composting and also involved to higher GHGs emission. Although, various additives such as biochar, wood ash, fly ash, kaoline, zeolite, lime, phosphogypsum, struvite salt and Ca-bentonite (Fang and Wong, 1999; Wong et al., 2009; Li et al., 2012; Luo et al., 2013; Wang et al., 2016; Awasthi et al., 2017a), and bulking agents like higher dosage of biochar, wood dust and agricultural wastes (Ermolaev et al., 2014; Santos et al., 2016) have been widely used for the mitigation of CH4 and N2O emission, and nitrogen conservation during the various kind of organic waste composting, but knowledge concerning about medical stone (MS) amendment into SS composting is limited. Therefore, it is hypothesized that MS addition into biosolids composting may have many benefits. Because micro porous-structure of MS has leads to provide optimum aeration for rapid biodegradation of OM and that could not only enhanced the microbial activities but act as biofilter to reduce the gaseous emission and nitrogen losses. Beside this, MS amendment adequate to buffer the pH and reduced the mobility and bioavailability of heavy metals during the bio-oxidative phase of composting (Wang et al., 2016). At present, from an extensive literature was searched and conclude that no earlier investigation described about carbon and nitrogen conservation through mitigation of CH4 and N2O emission and its relation of microbial biomass of carbon and nitrogen for biosolids composting employing MS. Hence, it is interesting to examine the feasibility of MS amendment for mitigation of GHGs and NH3 emission during the biosolids composting and resulting maturity indexes as well as the evolution of gene copy numbers of total bacteria (TB), ammonia oxidizing bacteria (AOB), total archaea bacteria (TAB) and ammonia oxidizing archaea bacteria (AOAB) microorganism involved in this process.

2.3. Gaseous emission and compost analysis The gaseous emission (NH3, CO2, CH4 and N2O) were determined according to our earlier study Awasthi et al. (2016a). The pH, EC, moisture content, TOM, TKN and TOC were analyzed as per TMECC (2002). The compost microbial biomass of carbon (MBC) and nitrogen (MBN) were quantified employing the chloroform fumigation-incubation and extraction method (Jenkinson and Powlson, 1976). 2.4. Molecular microbiology sampling and analysis Fresh compost samples collected on day 0, 3, 7, 14, 21, 28, 35 and 42 were used to extract DNA using a DNA extraction kit (Fast DNA SPIN Kit for Soil, Omega Biotek, Inc.) as per manufacturer’s directions. The extracted DNA was quantified using a spectrophotometer (Nano Drop 2000, Thermo, Japan), while DNA concentration and purity were observed on 1% agarose gels. Based on the concentration, DNA was diluted to 1.0 ng/μL using Milli-Q water. The PCR was performed using special bacterial universal primers 515F (5′GTGCCAGCMGCCGCGGTAAT-3′) and 806R (5′-GGACTACHVGGG TWTCTAA-3′), and a GC-clamp. All PCR reactions were carried out with Phusion® High-Fidelity PCR Master Mix (New England Biolabs). Touchdown PCR was used to amplify bacterial and archaea 16S V3 variable regions as follows; the PCR mixture (final volume 50 μL) contained 20 μL Premix Ex Taq (Takara Biotechnology), 0.4 μL of each primer (10 μM), 4 μL of five-fold diluted template DNA (1–10 ng) and 25.2 μL-sterilized water. Polymerase chain reaction (PCR) was performed according to Sun et al. (2016), while finally PCR products were purified using Qiagen Gel Extraction Kit and PCR Clean-up system (Qiagen, Germany). The concentrations of DNA in the PCR products were fluorometrically estimated employing the Qubit dsDNA HS Assay Kit (Invitrogen, Carlsbad, CA, USA), while the sequences were identified using by the help of Novogene (Miseq platform, Illumina, San Diego, CA, USA), Beijing, China. The total abundance of the groups of selected bacteria was analyzed according to Wang et al. (2017). The target gene data was revealed as mean slop of the gene copy number per 1.0 kg of the total feed stock (copies·kg−1-TS).

2. Materials and methods 2.1. Raw materials collection and processing The biosolids and wheat straw (WS) used in this study were collected from the Yangling sewage water treatment plant and Northwest A & F University, China. The MS was used as additive and purchase from Shijiazhuang Jiacheng Building Materials Co. Ltd., China. The biosolids and WS (bulking agent) were mixed at 1:1 ratio (dry weight basis) to achieved the ∼55% moisture content and ∼25 C/N ratio, while 1.0 kg of plastic spheres (non-biodegradable) were also mixed with initial feedstock to adjust the 0.5 kg/L bulk density. The physicochemical characteristics of raw materials are already provide in our previous study (Awasthi et al., 2017a), while MS pH (8.73 ± 0.02), moisture content (1.02 ± 0.04), and total organic matter (TOM), total organic carbon (TOC) and total Kjeldahl nitrogen (TKN) are not detectable.

2.5. Statistical analysis The average value of three replicates of each analysis were reported and the data were deal on the basis of two-way analysis of variance (ANOVA), while multiple comparison tests were also conducted to identified the least significance difference at p = 0.05 values employing SPSS v.21 software package for windows. The redundancy analysis (RDA) was implemented to identify the correlation of among the all physiochemical properties, gaseous emission and biological parameters during the biosolids composting by the using of Canoco 5 software.

2.2. Experiment design and compost sample collection The composting experiment was carried out in a series of 130-L invessel reactors. Details of the setup and operation of the composting process were provided in our previous report (Awasthi et al., 2017b). Five treatments were performed to assess the impact of MS-2%, 4%, 6% and 10% (on biosolids dry weight basis) for mitigation of gaseous emission and production of matured compost. The biosolids without any MS blended was used as control for comparison purpose. The ∼100-L of composting mixture was then loaded into the each composter for the present experiment, as described above. Prior to feedstock loading, the biosolids residue and WS were manually mixed thoroughly combined with additive to ensure the homogeneity. Next, the composting mass from the reactor was taken out and mixed

3. Results and discussion 3.1. Effect of medical stone amendment on maturity indexes The temperature is one of the important criteria to assess the overall composting process, at which plenty of the microbiological bio-oxidative reactions take place during composting. From a biological flyspeck, three composting passed through three main phases: first temperatures 971

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Fig. 1. Evaluation of temperature (a), pH (b), NH4+-N (c) and C/N ratio (d) profile of different treatments during biosolids composting. B + WS: biosolids + wheat straw (Control); B + WS + 2% MS: biosolids + wheat straw + 2% medical stone; B + WS + 4% MS: biosolids + wheat straw + 4% medical stone; B + WS + 6% MS: biosolids + wheat straw + 6% medical stone; B + WS + 10% MS biosolids + wheat straw + 10%medical stone; Results are the mean of three replicates and error bars indicates standard deviation.

treatments, which indicated the potential effect of the higher dosage of MS used in those mixture, respectively. The shorter bio-oxidative phase was also observed in previous studies of SS composting applied with different kind of additives (Villasenor et al., 2011; Awasthi et al., 2016b, 2017b). Treatments amended with 2% MS and 4% MS, the temperatures were decreased from the day 14 and then entered to maturation phase which could be due to continues abatement of bioavailable nutrients. The temperature profile of current study was in line with the previous findings by Awasthi et al. (2016b) and Wang et al. (2016), who investigated the effect of biochar and MS amendments on SS and PM composting and Sanchez-Garcia et al. (2015), who conducted composting experiment to identify the influence of biochar amendment during poultry manure composting for N mineralization and conservation. In general, microbial activities were directly influenced by the pH variation, because 6.5–7.5 pH is proposed by numbers of researchers for the active composting (Wong et al., 2009; Czekala et al., 2016). The changes of pH profile among the all five compost mixes during the composting are indicated in Fig. 1b. It was noticed that the pH was very fluctuating and ANOVA single factor test analysis also showed considerable variation among the all treatments (p < 0.05). At the start of the composting, the average pH of among the all feed stock mixture was 7.42 ± 0.07–8.06 ± 0.03. As illustrated in Fig. 1b, the pH of among the all MS applied treatments were sharply reduced from 8.06 ± 0.03 to 7.35 ± 0.034 (10% MS), 7.65 ± 0.06 to 6.86 ± 0.03 (6% MS), 7.60 ± 0.02 to 6.89 ± 0.08 (4% MS) and 7.50 ± 0.05 to 6.02 ± 0.036 (2% MS) and then constantly increased at the end of composting, which was attributed due to the augmentation of organic

above 55 °C to escalate the sanitization of pathogenic microbes and weed seeds (thermophilic phase), between 45 and 55 °C to enhance the microbial diversity and mineralization rate (first mesophilic phase) and then between 35 and 40 °C (maturation phase) (Amlinger et al., 2008; Sanchez-Garcia et al., 2015; Awasthi et al., 2017a). The evolution pattern of the temperature was varied among the all treatments (Fig. 1a). From the beginning of composting, the temperature sharply increased in 10% MS and 6% MS applied treatments and reached values more than 55 °C (thermophilic range) within 3 days and continue above this level for 5–7 days which is direct indicator of active microbial activities and organic matter mineralization. But control treatment temperature was never reached 50 °C through the composting period might be due to low microbial activities and buffering ability. The higher temperature was observed in 10% MS and 6% MS amended treatments than among the all treatments attributed due to the higher dosage of MS addition could increased porosity and alleviate acidity of composting mass during the bio-oxidative stage. Also, the first turning enhanced the aerobic microbial activities and composting in 10% MS and 6% MS amended treatments as reflected in the temperature profile. The reactivation of the microbial activities and organic waste mineralization by the addition of higher dosage of additives and turning was also observed in our earlier study Awasthi et al. (2016a) during the composting of SS amended with zeolite. Treatments applied with 2% MS and 4% MS were not showed any buffering effect (Fig. 1b), and then these treatments did not reach more than 50 °C temperature throughout the composting (Fig. 1a). However, the treatments applied with 10% MS and 6% MS, showed thermophilic conditions for prolonged period of time than among the other MS amended 972

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Chan et al., 2016). It was many authors confirmed that the mineralization of TOC and TOM could be possible region to reduce the C/N value and start to higher CO2 emission (Hao et al., 2001; Dias et al., 2010). Beside this, because of rapid TOM degradation, mass losses might assist to the raise of total nitrogen content which induce the decline of C/N ratio (Wang et al., 2016; Awasthi et al., 2017a). In present study, C/N ratios rapidly rose during the first 7 days in 6% and 10% MS applied treatments, and then gradually declined along with the composting processes (Fig. 1d). In contrast, the increasing trend of C/N ratios in 2 and 4% MS amended was late and occurred on day 10, while in control the rising trend of C/N ratio was observed until day 42. The percentage of C/N ratio recession in 10% MS was considerably higher than that in 6% MS applied treatment because of the active biological activities and substantial mineralization of TOM (Luo et al., 2013; Awasthi et al., 2016b, 2017a). At the end of biosolids composting, the C/N ratio was significantly higher in control (29.78), 2% MS (18.94) and 4% MS (19.68) applied treatments than that in 6% MS (16.62) and 10% MS (16.90), respectively, since that higher dosage of MS (6% and 10%) enhanced the TOC degradation and diminished the nitrogen losses. The final C/N ratios in 6% and 10% MS applied treatments comply within the standard value for a successful composting and its utilization for organic farming purposes (< 25) (Sommer and Moller, 2000; Tsutsui et al., 2013; Chan et al., 2016). The rising trend of C/N ratio at the beginning of biosolids composting were attributed due to the considerable amount of TN loss via NH3 emissions. However, the C/ N ratios declined among the all MS applied treatments afterwards, which was due to the mineralization of TOM and the conversion of complex nitrogen (protein, nucleic and ammonia acids) to organic nitrogen.

acid by the rapid degradation of organic matter. In addition, the consecutive TOC and total nitrogen degradation by microbial activities, and the carbonate formation was decreased composting mass pH, but as volatile organic compounds were break down and exoneration of NH4+-N and volatile ammonia, which could be increases pH value once again. Hence, the pH raised and arrived to 8.24 ± 0.05 (10% MS), 7.83 ± 0.02 (6% MS) and 7.66 ± 0.03 (4% MS) at the end of composting. However, Wang et al. (2016) examined that the 10% MS amendment was significantly buffer the pH during the PM composting, and our earlier finding with 10% zeolite applied gelatin industry sludge composting (Awasthi et al., 2016a) also confirmed that higher dosage of additives amended showed considerably higher pH than control. In contrast, 2% MS cannot considerably buffer the composting mass and also not provide the suitable pH 6.5–7.5 for active microbial activities during the bio-oxidative phase of composting (Fig. 1b), while for control treatments composting mass pH was initially observed optimum range, but does not appearance any pH increment with the composting process. Consequently, the decreasing trend of control treatment pH from the beginning of SS composting was also observed by many scientist (Villasenor et al., 2011; Ermolaev et al., 2014; Awasthi et al., 2016a), but this tendency of pH would prolonged on day 35 and then marginally elevated. This decline value of pH could be due to the augmentation of organic acids, which prohibit the biological activities and thus TOM mineralization. Such impact was also observed in control treatment temperature profile (Figs. 1a and 2a). However, analogous results obtained by many researchers (Villasenor et al., 2011; Li et al., 2012) during the SS and PM composting applied with zeolite and Cabentonite, Wang et al. (2016), who first time identify the buffering effect of MS for PM composting. Finally, the result showed that 6 and 10% MS applied treatments composts pH was within the range of standard (pH 7.5–8.5) for compost application (TMECC, 2002), but control was not matured at the end of composting. However, the 2–4% MS blended treatments compost pH was up to the standard limit but according to other analyzed parameters, this compost was not completely matured. Whilst this higher pH among the 2–4% MS blended treatments compost could be attributed due to biomass loss and precipitation of ammonium ions. The variation of NH4+-N is a crucial intimation of NH3 emission and nitrogen evolution. As shown in Fig. 1c, the NH4+-N concentrations among the all MS (4%, 6% and 10%) amended treatments increased rapidly from 358.96 to 3894.13 mg/kg DW (4% MS), 327.13 to 4160.19 mg/kg DW (6% MS) and 298.46 to 4341.16 mg/kg DW (10% MS) compost on day 7. In Control and 2% MS amended treatments, the rapid increasing trend was slightly late such as in 2% MS applied treatment from day 21 gradually decreased until to end the composting, while in control treatment increasing trend of NH4+-N was examined on day 35 and then slightly decreased. Compared to the initial value of NH4+-N, it increased by 12–14 times in MS amended treatments (4%, 6% and 10%), which was considerably higher than control and 2% MS applied treatments. The increasing trend of NH4+-N among the all MS (4%, 6% and 10%) applied treatments attributed due to the active decomposition of organic nitrogen to NH4+-N by the ammonification step in nitrogen cycle. Then, NH4+-N concentrations initiate to decline might be because of the assimilation as proteins, nucleic and amino acids by nitrifying and denitrifying bacteria (Fukumoto et al., 2003; Chan et al., 2016; Awasthi et al., 2017a). Previous researcher have been examined that decrease of NH4+-N could be used as a vital criteria in terms of an efficient composting and maturation of end product (Wong et al., 2009; Chowdhury et al., 2014; Jiang et al., 2016; Awasthi et al., 2016c). At the end of biosolids composting, NH4+-N concentration in 6% and 10% MS applied treatments was 446.34 mg/kg DW and 398.14 mg/kg DW compost, which was considerably lower than that observed among the all treatments (Fig. 1c). The reducing trend of C/N ratio is indicator of rapid decomposition of TOM and TOC, which is normally used as a important parameter to identify the compost maturity (Manios et al., 2007; Wong et al., 2009;

3.2. Effect of medical stone amendment on gaseous emission profile The evolution of biotic (CO2 and NH3 emissions) and abiotic (CH4 and N2O) activities in present biosolids composting for respective trial in premises of kg/gm of treated organic waste is illustrated in Fig. 2. 3.2.1. Changes of carbon dioxide and methane emission profile At the bio-oxidative phase of the composting, the CO2 production is inevitable due to active microbial activities and its emissions rapidly increased as the composting progress (first 1–10 days), and then steadily declined (14–28 days). From the beginning of composting, the maximum CO2 emission was observed in 10% MS and 6% MS applied treatments on day 3 and then the constantly reduced until the end of composting which showed the stability of the end product. The lowest CO2 emissions was observed from control treatment, possibly due to less buffering ability and unfavorable condition for rapid biological activities, and thus slow mineralization of organic waste occur. Meanwhile, maximum CH4 emissions were examined from the control treatment during the thermophilic phase of composting, which clearly indicated to be presence of anaerobic pockets and accumulation of volatile fatty acids, while this inhibitory impact was also observed in the temperature, pH and NH4+-N figures for control treatment (Fig. 1) (Wong et al., 2009; Shao et al., 2014; Jiang et al., 2016). In contrast, the 2% MS and 4% MS amended treatments were showed considerably low CO2 than 6% MS and 10% MS applied treatments (Fig. 2), which might be due to less adequate buffering ability of applied dosages of MS and then low aerobic microbial activities. Normally, CH4 is produced by the deoxidization of carbon dioxide/ hydrogen and CH3COOH by methanogenaic bacteria under oxygen deficiency conditions. In the present study, the CH4 emissions generally resulted during the early composting days (first 10 days) of control treatment, which is similar with previous findings (Manios et al., 2007; Dias et al., 2010; Yang et al., 2013; Awasthi et al., 2017a) for various kind of organic waste composting. Beside this, in all MS applied treatments were CH4 produced during the curing phase, which might be attributed due to the rapid mineralization of TOM during the early 973

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Fig. 2. Evaluation of CO2 (a), CH4 (b), NH3 (c) and N2O (d) profile of different treatments during biosolids composting. B + WS: biosolids + wheat straw (Control); B + WS + 2% MS: biosolids + wheat straw + 2% medical stone; B + WS + 4% MS: biosolids + wheat straw + 4% medical stone; B + WS + 6% MS: biosolids + wheat straw + 6% medical stone; B + WS + 10% MS biosolids + wheat straw + 10%medical stone; results are the mean of three replicates and error bars indicates standard deviation.

2014; Sanchez-Garcia et al., 2015), but were considerably lower than the earlier result observed in our previous study Awasthi et al. (2017b), in which ∼80 g/day CO2 emission was lost in the early phase of SS composting applied with a 12–18% wheat straw biochar. Hence, our results confirmed that 10% MS addition was much more effective to enhance the OM mineralization with significant (P < 0.001) high CO2 and less CH4 emission compare to other MS applied and control treatments.

phase and composting mass settlement, an anaerobic pockets is formed, thus considerable amount of CH4 is emitted from the all MS applied treatments. The CH4 emission from the control treatment reached its peak value on day 6 at 6.43 g CH4-C/kg/day, and from MS applied treatments significantly very low emission was observed through the composting process. The CH4 emission profile was ultimately persistent with that of temperature and CO2 emission for all treatments. However, the low MS dosages (2 and 4%) treatment had considerably the higher CH4 emission than 6 and 10% MS applied treatments, this was possibly because of the low MS dosage cannot maintain satisfactory oxygen concentration for rapid aerobic bacterial activities, and thus the oxygen deficiency or anaerobic condition conditions develop in the composting matrix. From our present results, we can understand that higher peak of CO2 emissions among the all MS blended treatments were possibly due to MS amendment proliferated the aeration within the composting mass and accelerate the mineralization of TOC, while without MS applied or control treatment composting mass cannot favor the optimal environmental condition according to microbial demand such as optimum pH (6.5–7.5) and aeration (0.2 L kg−1 DM min−1) (Guo et al., 2012). Consequently, the statistical analysis report also indicated that there was considerable difference among the 6% MS and 10% MS amended treatments CO2 and CH4 emission profile; while major asymmetry was observed between 2% MS, 4% MS blended and control treatments. In addition, our RD analysis was also indicated homologous correlation of CO2 and CH4 emission profile with temperature, pH and NH4+-N variation (Fig. 1). The CH4 and CO2 emission profiles were homogenous with many already published findings (Luo et al., 2013; Malinska et al.,

3.2.2. Nitrogen conservation through reduction of ammonia and nitrous oxide emission Profiles of NH3 and N2O emission from the composting treatments are shown in Fig. 2. The NH3 emission is sharply increased at the beginning with increasing temperature on day 1–7 in the 6% MS and 10% MS applied treatments, which was normally occurred by the rapid aerobic degradation or ammonification of organic nitrogen compounds (Amlinger et al., 2008; Tsutsui et al., 2013; Chan et al., 2016). In contrast, significantly very low NH3 emission was observed in control and 2% MS applied treatments through as compare to other MS amended treatments (Fig. 2a), might be due to high moisture and accumulation of organic acids could not favored the aerobic biological activities and thus decreased the OM mineralization. However, after first turning considerable amount of NH3 emission was observed in the control and 2% MS applied treatment, and that could be considered as significant amount of nitrogen loss. Among the all MS blended treatments, maximum NH3 was emitted on day 2 (6% MS) and day 3 (10% MS and 4% MS), and then sharply decreased with the fall in temperature from days 4–7 (Fig. 1a), while considerable amount of NH3 974

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emission rate was observed on day 5 from 2% MS applied treatment might be due to its low temperature (Fig. 1a). The results indicated that 10% MS addition was effective dosage to reduce the 10–22% TKN loss than other MS applied treatments and enhanced the aerobic microbial activities as well as rate of composting (Fig. 2c). Consequently, the statistical analysis also confirmed that the 10% MS amendment had a significant impact to reduced the NH3 emissions (P = 0.05) than 6% MS and 4% MS blended treatments, while no considerable differences were noticed among the other MS applied treatments composting process. The NH3 emission profiles from our study were uniform with the earlier published literature, and the basic concept for the profile change has been well described (Li et al., 2013; Awasthi et al., 2017a; Wang et al., 2017). The N2O emission profile is heterogeneous nature because it can be produced by two kinds of pathways such as denitrification/incomplete nitrification (Beck-Friis et al., 2000; Szanto et al., 2007; SanchezMonedero et al., 2010; Wang et al., 2016). In present study, the N2O emission was began from day 1 in among the all treatments; while relatively very high N2O emission was examined from the control and 2% MS blended composter (Fig. 2d). Compared to control and 2% MS blended treatments, the 4, 6 and 10% MS applied treatment showed lower N2O emission through the composting process. However, the initial N2O production among the all treatments was possibly due to the nitrification/denitrification processes that usually convert NH4+-N into N2O and NO, when the raw materials were stored and then some anaerobic pockets formed at the beginning of composting (Dias et al., 2010; Yang et al., 2013). The similar N2O emission profile was observed by previous scientist (El Kader et al., 2007; Maulini-Duran et al., 2014; Awasthi et al., 2016b; Jiang et al., 2016) for farm manure, organic fraction of municipal solid waste and SS composting. The maximum N2O emission was noticed in the 4, 6 and 10% MS applied composting mass from day 25 to 30 (Fig. 2d) and then steadily decreased; this investigation precisely proved the existence of anaerobic sites during the curing phase of composting, which could be due to the result of composting feed stock disposition decreased the O2 availability. These anaerobic sites create despite the constrained aeration and turning events supplied into composting reactors to eliminate the development of anaerobic condition in the composting mass. Among the all MS applied treatments, very low N2O emission was determined from the 10% MS blended treatment, which could be due to the rapid decomposition of organic matter and low NO3 concentration during the thermophilic phase (data not showed) as well as dominant NH4+-N concentration (Fig. 1e). Consequently, our findings clearly indicated that N2O emissions seemed to be also hampered for a considerable amount of total nitrogen losses from the all treatments. Similar results were also obtained in our previous study (Awasthi et al., 2016c, 2017a), where higher (8–18%) and lower (2–6%) dosage of wheat straw biochar mixed with initial composting mass to regulate the N2O emissions and total nitrogen losses during the biosolids composting process. Beside this, one-way analysis of variance also showed a significant (P < 0.05) difference among the all MS applied treatments. However, the LSD-t did not indicate considerable difference between the 6% MS and 10% MS added treatments, but a relatively high difference was observed than the control treatment. Furthermore, RDA clearly indicated that N2O emissions has significant correlation with CO2, NH3 and CH4 emission profile as well as temperature and NH4+-N transformation (Fig. 5).

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Fig. 3. Evaluation of microbial biomass of carbon (a), microbial biomass of nitrogen (b) profile during biosolids composting. B + WS: biosolids + wheat straw (Control); B + WS + 2% MS: biosolids + wheat straw + 2% medical stone; B+WS+ 4% MS: biosolids + wheat straw + 4% medical stone; B + WS+ 6% MS: biosolids + wheat straw + 6% medical stone; B + WS+ 10% MS biosolids + wheat straw + 10% medical stone; Results are the mean of three replicates and error bars indicates standard deviation.

differences among the both treatments. In contrast, the control and 2% MS applied treatments showed lowest content of MBC but significantly higher MBN content was observed in 6% and 10% MS blended composting mass than control and 4% MS amended treatments. In contrast, very low MBC (128.63 mg/kg) and MBN (15.43 mg/kg) contents were observed in control treatment. Similar MBC and MBN profiles were also observed in our previous study (Awasthi et al., 2017a), when Ca-bentonite combined with 12% biochar was applied for SS composting. However, lot of research has been done in composting but MBC and MBN profiles were not described which is very essential to know the mechanism of GHGs and NH3 emissions as well as its correlation with physicochemical profiles. The RDA results showed that MBC and MBN contents not have significant correlation with gaseous (CO2, CH4, NH3 and N2O) emission, temperature and microbial profile of SS composting (Fig. 5), might be due nutrient concentration and MS dosage variation.

3.3. Effect of medical stone amendment on microbial biomass The MBC and MBN profile of the different treatments composting feed stocks are presented in Fig. 3. The results indicated that MBC and MBN contents were gradually increased with the composting period and maximum content observed among the all treatments at the end of composting (Fig. 3). However, the MBC and MBN contents profile of 4% MS and 10% MS blended treatments were similar but considerable

3.4. Evolution of gene copy numbers The change of groups of bacteria indicates the overall process of ammonification and nitrification during the composting. In present study, these bacterial abundances were quantified by qPCR. The gene 975

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Fig. 4. Changes in the gene copy numbers of total bacteria (a), ammonia oxidizing bacteria (b), total archaea bacteria (c) and ammonia oxidizing archaea bacteria (d) during biosolids composting. B + WS: biosolids + wheat straw (Control); B + WS+ 2% MS: biosolids + wheat straw + 2% medical stone; B + WS+ 4% MS: biosolids + wheat straw + 4% medical stone; B + WS+ 6% MS: biosolids + wheat straw + 6% medical stone; B + WS+ 10% MS biosolids + wheat straw + 10% medical stone; Results are the mean of three replicates and error bars indicates standard deviation.

showed) and decreased of NH4+-N concentration (Fig. 1c). In addition, the profile of AOB was homogenous with the change of TB copy numbers, while the ratio of AOB/TB ranged from ∼0.0001% in initial feedstock to around ∼0.1% in the compost of each treatments. The number of AOAB was about 80-fold higher than the AOB in the raw material. Similar to the AOB, the abundance of AOAB gradually decreased to 8.4 × 109 copies·kg−1-TS on day 21 in all MS blended treatments and then dramatically increased until the end of composting. But in control treatment, the copy number of AOAB was decreased on day 28 and then steadily increased until day 42. Among the all MS applied treatments, the copy number of AOA was relatively higher than that of AOB during the first 21 days, while AOB copy number in control treatment first 14 day increased and then decreasing trend until the end of composting, which were also reflect in NH3 and NO2 emission profile (Fig. 2b and c). It seems like that AOAB could sustain under higher temperatures and NH4+-N concentrations, which are usually ominous for the growth of AOB (Fukumoto et al., 2003; Zeng et al., 2012; Li et al., 2013). Since significant amounts of nitrate-formed during the first two weeks (data not showed), therefore, AOAB was conceivably intended for the biological nitrification during the maturation phase of the biosolids composting. Moreover, among the all MS applied treatments, the gene copy numbers of AOAB decreased to 1.6 × 1011 copies·kg−1-TS on day 28, and then constantly elevated until the end of the biosolids composting. From the results, it was conclude that the numbers of AOAB were relatively the same as the copy numbers of AOB among the 6% and 10% MS blended treatments during the maturation phase of the composting. But considerable difference were observed between control and other (2% and 4% MS) treatments. Taking into explanation with the evidence that other groups of bacteria and archaea multiply or refrained at the carbon-copy way but it’s influenced by O2 availability and temperature profile, while the

copy numbers of TB and TAB are illustrated in Fig. 4a and c, and the amoA gene copy numbers of AOB and AOAB are provided in Fig. 4b and d. Among the all treatments, the copy numbers for TB were increasing trends throughout the composting, within the ranged between 1011 and 1013 copies·kg−1-TS. In contrast, in 4% MS applied treatment slightly TB gene abundance decreased from an initial 5.8 × 1013–5.2 × 1011 copies·kg−1-TS on day 14, and then resume and gradually increased until day 42 of the SS composting. The initial numbers of TA were around 2.3 × 1012–2.8 × 1014 copies·kg−1-TS, but this decreased to 1.8 × 1013–1.5 × 1012 copies·kg−1-TS with increasing MS dosages. But TA abundance was gradually decreased among the all treatments with the composting progress Fig. 4c. In the ubiquity of optimum aeration or O2, nitrifying bacteria oxidize parts of ammonium/ammonia to nitrate and then nitrite via biological nitrification. The NH3 oxidizing bacterial including AOB and AOAB are adequate to oxidized ammonium to nitrite. Meanwhile, the generated nitrous oxide is subsequently oxidized to nitrate by nitrite-oxidizing bacteria. Yao et al. (2011) has reported that NH4+ oxidation is the complex process in nitrification, where many metabolic/catabolic reactions occurred. Both AOB and AOAB were identified in the initial feedstock, and among the all treatments the initial copy number between 8.9 × 109 and 6.8 × 1013 copies·kg−1-TS, respectively (Fig. 4b). The number of AOB decreased to around 2.1 × 106 copies·kg−1-TS in the first 21 days among the all MS applied treatments, and this normally attributed due to the growth of AOB was hamper by the higher temperatures and lower pH (Fig. 1a and b). Subsequently, among the MS blended treatments, AOB gradually increased to 2.3 × 06 copies·kg−1-TS from the day 28 to until the end of composting process, and then higher abundance was noticed on day 42. Hence, the number of AOB re-produced ∼100-fold from day 21 to day 42. The increase in the AOB copy number from day 21 to 42 could be due to the increased nitrate concentration (data not 976

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Fig. 5. Redundancy analysis of gaseous emission and physicochemical properties during the biosolids composting. The correlation between gaseous emission, physical and nutrient transformation properties variables can be approximated by a perpendicular projection of the two different color variable arrow-tips onto the line overlaying the ordination axes. The length of the gaseous emission arrows is the multiple correlations of physiochemical properties with the ordination axes: Control (a), 2% medical stone (b), 4% medical stone (c), 6% medical stone (d) and 10% medical stone (e) applied biosolids composting.

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ratios of AOB/TB and AOAB/TA were used to identify the self-reliant dynamics of both groups of bacteria. The RDA of each treatments were indicated that the emission of NH3 and NO2 positively correlated with AOB and AOAB profile (Fig. 5). Therefore, it is believed that both AOB and AOAB were responsible for nitrification during the biosolids composting. 4. Conclusions The 10% MS amendment reduced the maximum GHGs (CH4 and N2O) emission and loss of TN. Considering these results, it was estimated that 6–10% MS could not only enhanced organic matter degradation and microbial biomass of nutrient (carbon and nitrogen) but also moderate the duration of the bio-oxidative phase and total composting duration by two weeks. The qPCR analysis of bacterial population confirmed that both AOB and AOAB were responsible for ammonia oxidation during the composting of biosolids. Overall, 10% MS addition for biosolids composting indicated to be an important environmental beneficial process for the reduction of GHGs emission and conservation of TOC and TN during biosolids composting. Acknowledgements The authors are grateful for the financial support from China Yangling Demonstration Zone Collaborative Innovation Major Projects on Production, Study, Research and Application. (No. 2016CXY-12) and China Postdoctoral Science Foundation (No. 2016 M602865). We are also thanks to our all laboratory colleagues and research staff members for their constructive advice and help. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.biortech.2017.09.201. References Amlinger, F., Peyr, S., Cuhls, C., 2008. Greenhouse gas emissions from composting and mechanical biological treatment. Waste Manage. Res. 26, 47–60. Awasthi, M.K., Wang, Q., Hongyu, C., Wang, M., Awasthi, S.K., Ren, X., Hanzhen, C., Li, R., Zhang, Z., 2017a. In-vessel co-composting of biosolid: focusing on mitigation of greenhouse gases emissions and nutrients conservation. Renew. Energy 1–10. http:// dx.doi.org/10.1016/j.renene.2017.02.068. Awasthi, M.K., Wang, Q., Huang, H., Rena, X., Lahori, A.H., Mahar, A., Ali, A., Feng, S., Li, R., Zhang, Z., 2016a. Influence of zeolite and lime as additives on greenhouse gas emissions and maturity evolution during sewage sludge composting. Bioresour. Technol. 216, 172–181. Awasthi, M.K., Wang, Q., Rena, X., Zhao, J., Huang, H., Awasthi, S.K., Lahori, A.H., Li, R., Zhang, Z., 2016b. Role of biochar amendment in mitigation of nitrogen loss and greenhouse gas emission during sewage sludge composting. Bioresour. Technol. 219, 270–280. Awasthi, M.K., Wang, Q., Huang, H., Li, R., Shen, F., Lahori, A.H., Wang, P., Guo, D., Guo, Z., Jiang, S., Zhang, Z., 2016c. Effect of biochar amendment on greenhouse gas emission and bio-availability of heavy metals during sewage sludge co-composting. J. Clean. Prod. 135, 829–835. Awasthi, M.K., Wang, M., Pandey, A., Chen, H., Awasthi, S.K., Wang, Q., Rena, X., Lahori, A.H., Li, D.S., Li, R., Zhang, Z., 2017b. Heterogeneity of zeolite combined with biochar properties as a function of sewage sludge composting and production of nutrient-rich compost. Waste Manage. 68, 760–773. Beck-Friis, B., Pell, M., Sonesson, U., Jonsson, H., Kirchmann, H., 2000. Formation and emission of N2O and CH4 from compost heaps of organic household waste. Environ. Monit. Assess. 62, 317–331. Bong, C.P.-C., Lim, Y.L., Ho, S.W., Lim, S.J., Klemes, J.J., Towprayoon, S., Ho, S.C., Lee, T.C., 2016. A review on the global warming potential of cleaner composting and mitigation strategies. J. Clean. Prod. 146 (10), 149–157. Cai, L., Chen, T.B., Gao, D., Yu, J., 2016. Bacterial communities and their association with the bio-drying of sewage sludge. Water Res. 90, 44–51. Chan, M.T., Selvam, A., Wong, J.W.C., 2016. Reducing nitrogen loss and salinity of ‘struvite” food waste composting by zeolite amendment. Bioresour. Technol. 200, 838–844. Chowdhury, M.A., Neergaard, A., Jensen, L.S., 2014. Potential of aeration flow rate and biochar addition to reduce greenhouse gas and ammonia emissions during manure

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2011. Links between ammonia oxidizer community structure, abundance, and nitrification potential in acidic soils. Appl. Environ. Microb. 77, 4618–4625. Zeng, Y., De Guardia, A., Ziebal, C., Junqueira De Macedo, F., Dabert, P., 2012. Nitrification and microbiological evolution during aerobic treatment of municipal solid wastes. Bioresour. Technol. 110, 144–152. Zeng, Y., Guardia De, A., Ziebal, C., De Macedo, F.J., Dabert, P., 2014. Nitrogen dynamic and microbiological evolution during aerobic treatment of digested sludge. Waste Biomass Valorization 5, 441–450.

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