Effective removal of methyl siloxane from water by sewage activated sludge microbes: biodegradation behavior and characteristics of microbial community

Bioresource Technology Reports 7 (2019) 100209

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Effective removal of methyl siloxane from water by sewage activated sludge microbes: biodegradation behavior and characteristics of microbial community

T

Yi Wanga,b,c, Zi-Feng Zhanga, Xi-Jun Xua, Chuan Chena, , Jia-Bao Xuc, Ling-Chao Konga, Peng Xiea, Chun-Miao Zhengb, Nan-Qi Rena, Duu-Jong Leed,e ⁎

a

State Key Laboratory of Urban Water Resource and Environment, School of Environment, Harbin Institute of Technology, Harbin 150090, China School of Environmental Science and Engineering, Southern University of Science and Technology, Shenzhen 518055, China Begbroke Science Park, Department of Engineering Science, University of Oxford, Woodstock Road, Oxford OX5 1PF, United Kingdom d Department of Chemical Engineering, National Taiwan University, Taipei 106, Taiwan e Department of Chemical Engineering, National Taiwan University of Science and Technology, Taipei 106, Taiwan b c

ARTICLE INFO

ABSTRACT

Keywords: Methyl siloxane Biodegradation Microbial community composition Adsorption

Current approaches of polydimethylsiloxanes (PDMS) treatment mostly rely on transferring from the liquid phase to other phases (gas and solid), leaving secondary pollution risks. In this study, biodegradation of PDMS in wastewater was found with activated sludge collected from real wastewater treatment plants (WWTPs). The highest biodegradation rates of dodecamethylpentasiloxane (L5) and tetradecamethylhexasiloxane (L6) reached up to 58.3% and 21.4% within 5-day anaerobic incubation at 37 °C under shaking at 100 rpm. Moreover, the microbial communities were investigated by using Illumina high-throughput sequencing of 16S-rRNA, which pointed Clostridium spp. as a potential candidate responsible for the biodegradation of PDMS under anaerobic condition. Notably, this is the first report of effective biodegradation of PDMS by microbes in activated sludge. This study could open up possibilities for bioavailable methods to depollute other emerging organic pollutants in water.

1. Introduction Intensive application of polydimethylsiloxanes (PDMS) in biomedicine, pharmaceuticals, and personal care products, has made them emerging pollutants in ecosystems and brought growing concerns. PDMS has been detected in a variety of environmental media including air, soil, sediment, river, and wastewater treatment plants (WWTPs) across the globe (Bletsou et al., 2013a; Genualdi et al., 2011; Krogseth et al., 2013; Lu et al., 2010). The bioaccumulation and toxicological effects of PDMS can potentially harm human health by causing adverse immune reactions, endocrine disruptions, and damaging connective tissues, livers and lungs (McGoldrick et al., 2014b), despite its low concentrations detected in natural water bodies (ng/L to μg/L) (Dekant and Klaunig, 2016; Jia et al., 2015; Lieberman et al., 1999; Stevens et al., 2001). European and Canadian regulators have listed several PDMS as priority pollutants in the aquatic environment and launched mandatory removal targets for WWTPs streams since 2008 (Wang et al., 2013b). Consequently, it is of much demand to develop feasible

⁎

solutions to eliminate these organic compounds from water. Several previous studies of the fates of PDMS in different environmental compartments are summarized (Table 1). Among them, the discharging effluent from WWTPs is a major source of PDMS loading into the aquatic environment. A number of studies have been conducted on PDMS in WWTPs, mainly focusing on the distribution of influent (0.06–227 μg/L), effluent (0.02–2.92 μg/L), biosolids (0.39–9.5 μg/g dw) and air (0.04–3 μg/m3). Volatilization and adsorption have been believed to be the main contributors to the removal of PDMS in WWTPs. However, none of the studies were conducted in a combined system of wastewater and sludge, nor estimated the role of microorganisms in degrading PDMS. Compared with volatilization and adsorption which transfer the pollutants to another environmental medium without any mass reduction, microbes can adsorb and assimilate the pollutants into biomass via biodegradation, leaving no secondary harm to the environment and human. To the best of our knowledge, no evidence has been found concerning functional microbes which can effectively degrade PDMS.

Corresponding author. E-mail address: [email protected] (C. Chen).

https://doi.org/10.1016/j.biteb.2019.100209 Received 17 March 2019; Received in revised form 25 April 2019; Accepted 26 April 2019 Available online 02 May 2019 2589-014X/ Crown Copyright © 2019 Published by Elsevier Ltd. All rights reserved.

Bioresource Technology Reports 7 (2019) 100209 (Bletsou et al., 2013a; Bletsou et al., 2013b; Capela et al., 2017; Parker et al., 1999; Sanchís et al., 2013; Wang et al., 2015; Wang et al., 2013b; Zhang et al., 2011)

(Batley et al., 1991; Borga et al., 2012; McGoldrick et al., 2014a; Stevens et al., 2001; Zhang et al., 2011) (Xu and Chandra, 1999; Xu et al., 1998; Zhang et al., 2011) (Genualdi et al., 2011; Krogseth et al., 2013; Lu et al., 2010; Pieri et al., 2013)

This is the first study which provides experimental evidence on biodegradation processes of PDMS and provides new insight to remove PDMS by activated sludge in WWTPs, which would be taken account into the future upgrade of treatment processes in WWTPs. Firstly, we evaluated the removal of PDMS in different phases in the activated sludge collected from a WWTP. We then estimated biodegradation rates of PDMS under various culturing conditions. Finally, by using highthroughput 16S rRNA sequencing, we mapped the structure of the microbial community and identified the possible functional bacteria in the activated sludge that are responsible for biodegradation. 2. Materials and methods 2.1. EPI Suite simulation EPI Suite was utilized to simulate PDMS removal and estimate the possibility of biodegradation of PDMS in solution. EPI Suite is a software developed by USEPA and Syracuse Research Corp (US EPA, 2016), aiming at assessing physical and chemical properties and environmental fate parameters, such as adsorption, volatilization, and biodegradation, when experimental values are not available. 2.2. Collection of activated sludge and wastewater All wastewater and activated sludge used in this study were collected from Wenchang Wastewater Treatment Plant, located in Harbin, Heilongjiang Province, China. The activated sludge samples were collected at three different locations from the secondary sedimentation tank and then mixed well. Wastewater samples were collected from the influent of the WWTP and instantly stored at 4 °C upon their arrival until utilization. The concentrations of main water quality parameters were listed in Table 2. Briefly, all samples were collected by the aluminum barrel and sealed with caps with Teflon gaskets and stored at 4 °C until use. The activated sludge was pre-acclimatized to culturing conditions (37 °C with shaking at 100 rpm) of growth for 7–15 days with wastewater collected from WWTP before the beginning of the experiments.

Volatilization, Adsorption, Hydrolysis Adsorption, Hydrolysis Photodegradation, Chemical reaction (with OH and NO3 radical and O3) Volatilization, Adsorption, Biodegradation

2.3. Batch experiments In order to simulate and reflect the biodegradation of PDMS which possibly occurs in WWTP, we use the real domestic sewage taken from the influent of WWTP, not the artificial wastewater in batch experiments. A total of 200 mL mixture of domestic sewage (50 mL) and activated sludge (150 mL) were poured into the bottle. A standard solution of PDMS200, 5cSt (Sigma-Aldrich) which contains 10 target methyl siloxanes, was added to each bottle until the concentration of PDMS reached 2 mg·L−1. The composition of PDMS200, 5cSt was determined by gas chromatography-flame ionization detection (GC-FID), described in an earlier study (Horii and Kannan, 2008). The biodegradation of methyl siloxane in wastewater by activated sludge was conducted under four conditions: aerobic incubation, anaerobic incubation, anaerobic incubation with 200 mg·L−1 KNO3 addition as nitrate and anaerobic incubation with 500 mg/L NaSO4 addition as sulfate. Each condition had a corresponding control group with an addition of 4% paraformaldehyde to inactivate the microorganisms in the sludge. Four replicates were performed in each group (32 serum bottles in total). The anaerobic experiments were conducted in 250 mL sealed glass bottles while aerobic experiments were conducted by using 250 mL Erlenmeyer flasks. Under all anaerobic conditions, the bottles were flushed with nitrogen gas for 10 min to remove the oxygen before incubation. Microorganisms in the bottles were cultured for a week at 37 °C with shaking at 100 rpm. Samples were harvested every 24 h in 10 mL amber glass bottles and stored at 4 °C until pretreatment and analysis. The background values of methyl siloxane may be high and interfered by ambient factors. To minimize interferences, any cosmetics or

/ + Wastewater treatment plant

16.7–315d 4.1–588d 10–27d − Aquatic Soil and sediment Atmosphere

References Fate Half-lives Biodegradation Compartment

Table 1 Summary of fates of PDMS in different environmental media (“+” indicated “with” or “presence”. “−” indicated “without” or “absence”. “/” indicated “unknown”).

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and Ci is the initial content of methyl siloxane. All of Cc, Ce, and Ci were measured.

Table 2 Concentrations of water quality parameters in the influent of Wenchang WWTP. Parameters

Concentrations (mg/L)

COD NH4+-N NO3−/NO2−-N TN TP

2.6. Analysis of microbial community structure by 16S rRNA sequencing

285 ± 18.5 25 ± 2.3 10 ± 1.9 35 ± 4.6 3.2 ± 1.0

Bioinformatic analysis by Illumina high-throughput sequencing of 16S rRNA was performed on activated sludge samples after 5-day incubation under aerobic condition, anaerobic condition, anaerobic condition with nitrate addition and anaerobic condition with sulfate addition, respectively. One milliliter of the activated sludge was taken for DNA extraction. Samples were centrifuged for 5 min at 13000 rpm at 4 °C, and the pellets were obtained for DNA isolation using the Fast DNA Spin Kit for Soil (MP Biomedicals) and the Fast Prep Instrument (MP Biomedicals) and quantified by a Qubit fluorimeter. PCR amplification using 16S primers 926Fw (AAACTYAAAKGAATTGRCGG) and 1392R (ACGGGCGGTGTGTRC), followed by 15 cycles with FLX titanium primers 454T_RA_X and 454T_FwB. High-throughput sequencing of the PCR products was performed by Illumina PE250. Analysis of operational taxonomic units (OTUs) was carried out by Mothur at a similar sequencing level of 97% (Schloss et al., 2009). OTUs were used for evaluation of community diversity. One of the naive Bayesian classifiers, RDP classifier, which can rapidly and accurately identify taxonomy, was used to analyze the composition and structure of microbial community with the Silva database (Quast et al., 2013).

Characteristics of activated sludge in Wenchang WWTP. Parameter

Value

Water content (%) pH VSS (mg/L) TSS (mg/L) SCOD (mg/L) SCOD/TCOD (%)

98.5 ± 0.1 6.8 ± 0.1 10,915 ± 365 19,719 ± 432 57 ± 5.5 0.249 ± 0.019

other products that may contain methyl siloxane were avoided during the experiment. Moreover, no plastic laboratory supplies were used, and glass apparatuses were washed with acetone and hexane solutions three times and dried under 60 °C before use.

3. Results and discussion

2.4. Gas Chromatography-Tandem Mass Spectrometry (GC–MS/MS) analysis

3.1. Biodegradation of methyl siloxane under different culturing conditions

We modified and optimized the detection method by Zhang et al. (Zhang et al., 2011), by utilizing an Agilent Technologies 7000GC–MS/ MS in MRM mode, which can quantify targeting methyl siloxanes by ion pairs, to improve accuracy and reduce the background noise of methyl siloxane. PDMS residue in sludge and in water was measured separately in all experiment and control sets. Samples were centrifuged for 5 min at 8000 rpm to separate liquid and solid components, both of which were pretreated by shaking for 30 min (sludge) or 5 min (liquid), and then washed three times with methylene chloride. Sludge samples were centrifuged for 10 min at 10000 rpm again and the solvent layer was transferred to round bottom flasks. Five milliliter of isooctane was added and concentrated to 2–3 mL by rotary evaporator with the water bath temperature at 31 °C. Afterward, the extract was transferred to a graduated tube and gently blown to 1 mL with nitrogen gas for GC–MS/ MS analysis. The extract samples were transferred to chromatography vials with the silica gel shim replaced with aluminum foil to prevent pollution. The analysis was done on an HP-5 column (30 m × 250 μm × 0.25 μm). Before the sample analysis, a series of different concentration standard solutions of methyl siloxane (10, 20, 50, 100, 200, 500, 1000 ng·mL−1) was used to calibrate the instrument; correlation coefficients of 0.9966 to 0.9995 were obtained. After several blank samples, 2 μL of the samples were injected in pulse splitless mode at 200 °C. The column oven temperatures were programmed from 40 °C to 220 °C in 9 min, from 220 °C to 280 °C at a rate of 60 °C·min-1, held for 3 min, and finally followed by 10 min of post running time at 300 °C.

Estimation Program Interface (EPI suite) was used to simulate the environmental fates of some common PDMS including cyclic and linear methyl siloxanes in wastewater. The simulation results indicated that the adsorption is the main fate, while biodegradation of all methyl siloxanes might occur at relatively low levels from 0.61% to 0.78% due to the chemical structure and their partition coefficients, such as Log Kow, Log Koc and Log Koa of PDMS (Table 3). Given this simulation, we performed experiments to validate the removal of these PDMS in real wastewater collected from a WWTP. Upon collection of activated sludge, detection of the added PDMS concentration in the liquid phase and in the solid phase was carried out by improving a previously described GC–MS/GC method by Zhang et al. (Zhang et al., 2011). The real content of the sample is the measured value of the experimental sample minus the value of blank sample. The recovery rate is the measured real content divided by the theoretical concentration added to the sample. This is able to reflect the influence of the target compound on the process of pretreatment. In our result, a high recovery rate ranged from 92% to 110% was achieved on all PDMS from D4 to D6 (cyclic PDMS) and L5 to L11 (linear PDMS) (Table 4). Activated sludge was then cultured in situ under different conditions including aerobic incubation, anaerobic incubation, anaerobic incubation with the addition of nitrate and anaerobic incubation with the addition of sulfate. In parallel, 4% paraformaldehyde was added to deactivate the microbial activity and calculate the degree of biodegradation (Hancock et al., 1982). Concentrations of D4-D6 and L5-L11 were measured on day 0, 1, 2, 3, 4 and 5 of culture, and biodegradation rates were calculated as the differences in PDMS residue between in the activated and the deactivated sludge. Among all PDMS, significant biological removal was only detected in removing dodecamethylpentasiloxane (L5) and tetradecamethylhexasiloxane (L6). Dynamic concentrations of L5 and L6 are presented in Fig. 1 (L5) and Fig. 2 (L6), and biodegradation rates are summarized in Table 5. The PDMS removed before day 1 in Figs. 1a, b and 2a, b was almost achieved through adsorption of activated sludge, with the extension of culturing time, the PDMS adsorbed gradually went down and the removal contributed by microorganisms went up until the PDMS

2.5. Calculation of biodegradation rate Biodegradation rate was estimated by calculating the differences in the overall PDMS between the experiment group with biological activity and the control group that has been treated with the antimicrobial agent. The calculation method is as follows:

Biodegradation Rate (%) = (Cc–Ce)/Ci × 100% where Cc is the total content of methyl siloxane in the control group without biological activity due to the addition with paraformaldehyde, Ce is the total content of methyl siloxane in the experimental group, 3

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Table 3 Fates of PDMS in wastewater simulated by EPI 4.1. Material

Abbreviation

Adsorption

Biodegradation

Residue in water

Evaporation

Octamethyl cyclotetrasiloxane Decamethyl cyclopentasiloxane Dodecamethyl cyclohexasiloxane Dodecamethylpentasiloxane Tetradecamethylhexasiloxane Hexadecamethylheptasiloxane Octadecamethyloctasiloxane Eicosamethylnonasiloxane Docosamethyldecasiloxane Tetracosamethylhendecasiloxane

D4 D5 D6 L5 L6 L7 L8 L9 L10 L11

83.53% 91.56% 93.20% 93.14% 87.35% 93.26% 93.26% 93.26% 93.26% 93.26%

0.61% 0.75% 0.78% 0.78% 0.67% 0.78% 0.78% 0.78% 0.78% 0.78%

15.25% 5.68% 5.96% 5.94% 4.01% 5.06% 5.06% 5.96% 5.96% 5.96%

0.61% 2.01% 0.07% 0.14% 7.06% 0.00% 0.00% 0.00% 0.00% 0.00%

adsorbed kept constant after day 3. On day 5 of the anaerobic culturing, significantly lower L5 was detected in the activated sludge compared with that in the deactivated sludge (p < .0001) (Fig. 1a). Furthermore, an analysis of residue L5 in different phases of either liquid or solid was done to understand the mechanism of transfer and degradation (Fig. 1a). In the deactivated sludge, a decrease in L5 concentration from day 0 to day 2 was observed in the liquid phase, accompanied by an increase in the solid phase. This is due to the low solubility of the PDMS in water thus an equilibrium achieved after day 2 between two phases. On the other hand, an earlier and more dramatic drop of the liquidphase L5 occurred on day 1 in the activated sludge, followed by a continuous reduction from day 1 to day 3. Notably, a decrease in the solid phase was also observed from day 1 to day 3. This suggests the additional mechanism of removal, possibly contributed by microbes, other than a physical transfer between two phases. The more reduced concentration observed in the liquid phase could be due to a formation of microbial flocs, which are microbes grouped together by extracellular polymeric substances (EPS) excretion, resulting in flocculation (Wilen et al., 2008). Activated sludge flocs have been proven to have surface “slime-layers” containing polysaccharides composed of neutral sugars and glucuronic acid which has strong capability to adsorb organic materials (Steiner et al., 1976). The biodegradation rate on day 5 under anaerobic condition was calculated to be 58.3% ± 11.5% for L5, which is the highest among all conditions (Table 5). Few studies focus on the biodegradation of PDMS, our data showed that the anaerobic condition benefits its biodegradation, which is in accordance with some studies showing that the removal of some refractory organic pollutants was the most efficient under anaerobic condition (Balamurugan et al., 2011). PDMS can be seen as persistent organic pollutants (Xu et al., 2012), anaerobic bacteria could break down these pollutants that resist aerobic degradation (Zheng et al., 2013). In order to highlight possible roles of nitrate-reducing and sulfatereducing bacteria in biodegradation of methyl siloxane, we sought to monitor the PDMS concentration with the addition of nitrate or sulfate (Fig. 1b and c). Nitrate and sulfate are two common contaminants in

wastewater (Yamashita and Yamamoto-Ikemoto, 2014) and it has been shown that the growth of substrate-removing bacteria is accelerated when supplemented with SO42− and NO3− under anaerobic condition (Sawada et al., 1978). Surprisingly, a decrease in the total content of L5 was observed in the deactivated sludge with the presence of nitrate, attributed to a significant reduction in liquid phase from day 0 to day 5 (Fig. 1b). This is different compared to all other anaerobic conditions in which equilibrium was reached between two phases, suggesting a chemical conversion by nitrate. Nitrate radical is a strong oxidant, reacting with a wide range of volatile organic compounds (VOCs) (Waring and Wells, 2015), PDMS is a class of VOCs (Wang et al., 2013a). It was observed that microbial degradation also played an important role in the total removal of PDMS. A more decreasing trend was observed in activated sludge with a difference calculated as 20.1 ± 9.2% (p < .0001) (Table 5). Interestingly, the biodegradation rate with an addition of nitrate was lower than that without nitrate. Chemical conversion of L5 was not seen in deactivated sludge with sulfate as indicated in Fig. 1c. A phase transfer from liquid to solid happened quickly from day 0 to day 1, followed by an equilibrium reached at later time points. The differences in L5 residue on day 0 between activated and deactivated sludge was calculated to be 10.0 ± 5.9% (p = .02), which is the lowest among all anaerobic conditions (Table 5). This suggests that the growth of sulfate-reducing bacteria may inhibit the growth of the bacteria that are responsible for degrading methyl siloxane, due to the competitions for the limited organic carbon source between two communities (Greene et al., 2003). Concentrations of L5 in aerobic condition were presented in Fig. 1d without the calculation of biodegradation rate because the degree of evaporation could not be taken into consideration in an open system. Interestingly, we observed differences in PDMS concentration in different environmental media with or without the presence of microbes. When the biological activity was present, we found an immediate transfer of PDMS from the liquid phase to the solid phase on day 1, followed by small fluctuations at later time points. Without biological activities, little medium transfer was found and the majority was removed in the water. This suggests a dominant role of adsorption (occupy 74% in total removal) by activated sludge was played because the high organic carbon adsorption coefficient, the log Koa of L5 is 8.101. On the other hand, volatilization contributed the most in the deactivated sludge (occupy 72% in total removal). The activated sludge can continue to adsorb L5 through the biodegradation, while L5 is preferred to volatile than adsorb by deactivated sludge in the control tests without biological activities due to its high log Kaw of 9.61 (Wang et al., 2013b). Concentrations of L6 were measured in the same conditions as of L5 (Fig. 2). The biodegradation rates on day 5 of anaerobic, anaerobic with nitrate and anaerobic with sulfate culturing were calculated to be 21.4 ± 6.3% (p < .001), 12.5 ± 6.3% (p < .001) and 0.9 ± 0.7% (p = .05), respectively (Fig. 2 and Table 5). Similar to L5, the highest biodegradation occurred during anaerobic culturing and lowest biodegradation was observed with an addition of sulfate. In aerobic

Table 4 The recovery rate of added PDMS by GC-MS/GC. Target material

Wastewater (water phase)

Sludge (sludge phase)

D3 D4 D5 D6 L5 L6 L7 L8 L9 L10 L11

93% 92% 97% 101% 95% 98% 99% 99% 100% 95% 97%

103% 105% 99% 104% 99% 95% 110% 105% 110% 98% 95%

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Fig. 1. Concentrations of L5 in different culturing conditions of (a) anaerobic (b) anaerobic with nitrate added (c) anaerobic with sulfate added and (d) aerobic. Standard deviation is drawn as error bar at each point from 4 replicates. Biodegradation: the total content of liquid and sludge in the control tests minus the total content of liquid and sludge in the experimental batch tests. Student t-test was used to indicate statistical differences between experiment group and control group at day 5, in which ns indicates non-significant, * indicates p < .05, ** indicates p < .01, *** indicates p < .001, **** indicates p < .0001.

culturing, 86% of the reduction of L6 was contributed by adsorption in activated sludge and 87% was contributed by volatilization when biological activity was absent (Fig. 2d). Compared with L5, L6 was more inclined to vaporize other than being absorbed, which is consistent with their physicochemical properties and the results simulated by EPI 4.1 (Table 3). The log Koa of L5 and L6 are 8.101 and 6.325 respectively, this ratio is octanol-air partition coefficient, which indicates the capacity of adsorption and evaporation. Notably, biodegradation of L6 in all conditions was lower than that of L5 (Figs. 1 and 2) while L7-L11 and D4-D6 showed no differences in residue concentrations between activated sludge and control (Fig. 3). Compared with D4 and D6, the concentration of D5 in the sludge decreased more due to the volatilization. According to the chemical structure and the partition coefficient, the percentage of volatilization of D5 is higher (Table 3). It is well-known that longer polymer chain length and more complicated conformations (cyclic rings) can increase the susceptibility of a polymer to a range of degradable agents e.g. various oxidative, chemicals and biological (Stivala, 1980).

sludge), AAS-N (anaerobic activated sludge with nitrate addition) and AAS-S (anaerobic activated sludge with sulfate addition), were analyzed by Illumina high-throughput sequencing of 16S rRNA to compare the abundance, diversity and composition of the bacterial community (Fig. 4, Table 6). A total of 483 genera were detected in all samples. IAS has the highest bacterial diversity with 445 species, while AAS-S has the lowest with 374 species. In IAS, Lactococcus was identified as the most dominant species, with an abundance of 21.8% (Fig. 4), which is frequently found in dairy products (Widyastuti et al., 2014). The next dominant two species were SJA-28 and Comamonadaceae, with an abundance of 5.1% and 5.0%, respectively. Compared with IAS, there was significantly more abundant Clostridium in AAS, increased from 1.1% to 13.0%, accompanied by a decrease in Lactococcus abundancy (15.0%) (Fig. 4). In addition, Azotobacter, Ruminococcaceae, Sedimentibacter, Lachnoclostridium, Anaerofilum, and Acidaminobacter were also enriched in AAS, suggesting a potential role of these bacteria in degrading PDMS. Similar with that in AAS, Lactococcus, Clostridium, and SJA-28 were also dominant among 412 genera in AAS-N with abundances of 16.7%, 9.8%, and 7.4%, respectively (Fig. 4). An enrichment of Clostridium was particularly observed from 1.1% to 9.8% compared with IAS. Except for Sedimentibacter, other five species that were enriched in AAS were also more abundantly existed in AAS-N after adding PDMS. On the other hand, the community structure in AAS-S changed

3.2. Dynamic of microbial community composition The initial activated sludge (IAS) and the activated sludge after 5day anaerobic incubation, represented as AAS (anaerobic activated 5

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Fig. 2. Concentrations of L6 in different culturing conditions of (a) anaerobic (b) anaerobic with nitrate added (c) anaerobic with sulfate added and (d) aerobic. Standard deviation is drawn as error bar at each point from 4 replicates. Biodegradation: the total content of liquid and sludge in the control tests minus the total content of liquid and sludge in the experimental batch tests. Student t-test was used to indicate statistical differences between experiment group and control group at day 5, in which ns indicates non-significant, * indicates p < .05, ** indicates p < .01, *** indicates p < .001, **** indicates p < .0001.

rates (58.3% and 20.1% of L5, 21.4% and 12.5% of L6). While most of the Clostridium spp. are strictly anaerobic, hydrogen-producing bacteria (Hallenbeck and Benemann, 2002), some species can degrade complex and recalcitrant organic compound, such as aromatic hydrocarbons (Arulazhagan and Vasudevan, 2009), trimethylamine, and cellulose (Collins et al., 1994). Our results suggest a similar degradation of methyl siloxane due to similar chemical structures of these organic compounds. Nevertheless, further functional analysis should be done to confirm Clostridium as a potential candidate in PDMS biodegradation.

Table 5 Biodegradation rates under different culturing conditions with the residence time of 5 days (“/” indicate “without”). Experiment conditions

Aerobic Anaerobic Anaerobic adding nitrate Anaerobic adding sulfate

Biodegradation rate (%) L5

L6

/ 58.3 ± 11.5 20.1 ± 9.2 10.0 ± 5.9

/ 21.4 ± 9.6 12.5 ± 6.3 0.9 ± 0.7

4. Conclusions

dramatically compared with the others. The primarily abundant genera in AAS-S were Acinetobacter, Acidaminobacter, Comamonadaceae, and Enterobacteriaceae, with an abundance of 10.3%, 7.0%, 6.8%, and 5.9%, respectively. Acinetobacter and Acidaminobacter are generally known to exist in the sulfate-reducing system in activated sludge and can greatly enhance phosphate removal (Hesselmann et al., 1999; Cloete and Steyn, 1988; Wagner et al., 1994). Other sulfate-reducing bacteria were also found to be selectively enriched due to the presence of SO42−, such as Thermomonas, Simplicispira. This is consistent with previous research showing an appearance of Simplicispira in the activated sludge after adding sulfate (Liu et al., 2005; Lu et al., 2007). Comparing the microbial community composition to biodegradation rate at different conditions, Clostridium was the most enriched in AAS and AAS-N, both of which correspond with the highest biodegradation

Besides adsorption and volatilization, we presented the first experimental evidence of microbial biodegradation in the effective removal of PDMS in wastewater. Among four different culturing conditions (5-day incubation at 37 °C), the highest biodegradation efficiency of L5 and L6 (8.3% and 21.4%) was obtained under the anaerobic condition. However, L7-L11 and D4-D6 were not biodegradable. High throughput sequencing of 16S rRNA indicated Clostridium may be a potential candidate for the biodegradation of PDMS. Our study fills the knowledge gap of the biodegradation of PDMS in wastewater and provides new insight into the transformation of some emerging pollutants, like PDMS.

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Fig. 3. Concentrations of residue L7-L11 and D4-D6 under anaerobic incubation.

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Fig. 4. Relative abundance of the major microbial type (> 1%) at the genus level in these four samples.

References

Table 6 Analysis of OTUs, abundance and diversity of microbial community in different culturing conditions. Samples

Sequences

OTUs

Chao

Shannon

51,015 47,661 48,417 38,723

1161 1096 1023 893

1209 1176 1133 1008

5.08 4.85 4.56 4.92

IAS AAS AAS (NO3−) AAS (SO42−)

Arulazhagan, P., Vasudevan, N.J.M.p.b, 2009. Role of a moderately halophilic bacterial consortium in the biodegradation of polyaromatic hydrocarbons. 58 (2), 256–262. Balamurugan, B., Thirumarimurugan, M., Kannadasan, T., 2011. Anaerobic degradation of textile dye bath effluent using Halomonas sp. Bioresour. Technol. 102 (10), 6365–6369. Batley, G., Hayes, J.J.M., Research, F., 1991. Polyorganosiloxanes (Silicones) in the Aquatic Environment of the Sydney Region. vol. 42(3). pp. 287–293. Bletsou, A.A., Asimakopoulos, A.G., Stasinakis, A.S., Thomaidis, N.S., Kannan, K., 2013a. Mass loading and fate of linear and cyclic siloxanes in a wastewater treatment plant in Greece. Environ. Sci. Technol. 47 (4), 1824–1832. Bletsou, A.A., Asimakopoulos, A.G., Stasinakis, A.S., Thomaidis, N.S., Kannan, K., 2013b. Mass loading and fate of linear and cyclic siloxanes in a wastewater treatment plant in Greece. Environ. Sci. Technol. 47 (4), 1824–1832. Borga, K., Fjeld, E., Kierkegaard, A., McLachlan, M.S., 2012. Food web accumulation of cyclic siloxanes in Lake Mjosa, Norway. Environ. Sci. Technol. 46 (11), 6347–6354. Capela, D., Ratola, N., Alves, A., Homem, V.J.E.i., 2017. Volatile methylsiloxanes through wastewater treatment plants–a review of levels and implications. 102, 9–29. Cloete, T.E., Steyn, P.L., 1988. A combined membrane filter-immunofluorescent technique for the in situ identification and enumeration of acinetobacter in activated sludge. Water Res. 22 (8), 961–969. Collins, M., Lawson, P., Willems, A., Cordoba, J., Fernandez-Garayzabal, J., Garcia, P., Cai, J., Hippe, H., Farrow, J.J.I.J.o.S., Microbiology, E., 1994. The phylogeny of the genus Clostridium: proposal of five new genera and eleven new species combinations. 44 (4), 812–826. Dekant, W., Klaunig, J.E., 2016. Toxicology of decamethylcyclopentasiloxane (D5). J.R.T., Pharmacology 74, S67–S76. Genualdi, S., Harner, T., Cheng, Y., Macleod, M., Hansen, K.M., van Egmond, R., Shoeib, M., Lee, S.C., 2011. Global distribution of linear and cyclic volatile methyl siloxanes in air. Environ. Sci. Technol. 45 (8), 3349–3354. Greene, E.A., Hubert, C., Nemati, M., Jenneman, G.E., Voordouw, G., 2003. Nitrite reductase activity of sulphate-reducing bacteria prevents their inhibition by nitratereducing, sulphide-oxidizing bacteria. Environ. Microbiol. 5 (7), 607–617. Hallenbeck, P.C., Benemann, J.R., 2002. Biological Hydrogen Production; Fundamentals and Limiting Processes. vols. 27(11–12). pp. 1185–1193. Hancock, W.W., Becker, G.J., Atkins, R.C., 1982. A comparison of fixatives and

Conflict of interest The authors declare no conflict of interest. Acknowledgements This research was supported by the National Natural Science Foundation of China (Grant No. 51576057 and 51676057), the National Key Research and Development Program (No. 2016YFC0401102-2), China Postdoctoral Science Foundation funded project (AUGA 4130903217), Heilongjiang Provincial Postdoctoral Science Foundation funded project (AUGA 4110002617) and the State Key Laboratory of Urban Water Resource and Environment (Harbin Institute of Technology) (No. 2015DX04). Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.biteb.2019.100209. 8

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Y. Wang, et al. immunohistochemical technics for use with monoclonal antibodies to cell surface antigens. Am. J. Clin. Pathol. 78 (6), 825–831. Hesselmann, R.P.X., Werlen, C., Hahn, D., van der Meer, J.R., Zehnder, A.J.B., 1999. Enrichment, phylogenetic analysis and detection of a bacterium that performs enhanced biological phosphate removal in activated sludge. Syst. Appl. Microbiol. 22 (3), 454–465. Horii, Y., Kannan, K., 2008. Survey of organosilicone compounds, including cyclic and linear siloxanes, in personal-care and household products. Arch. Environ. Contam. Toxicol. 55 (4), 701. Jia, H., Zhang, Z., Wang, C., Hong, W.J., Sun, Y., Li, Y.F., 2015. Trophic transfer of methyl siloxanes in the marine food web from coastal area of Northern China. Environ. Sci. Technol. 49 (5), 2833–2840. Krogseth, I.S., Kierkegaard, A., McLachlan, M.S., Breivik, K., Hansen, K.M., Schlabach, M., 2013. Occurrence and seasonality of cyclic volatile methyl siloxanes in Arctic air. Environ. Sci. Technol. 47 (1), 502–509. Lieberman, M.W., Lykissa, E.D., Barrios, R., Ou, C.N., Kala, G., Kala, S.V.J.E.h.p., 1999. Cyclosiloxanes produce fatal liver and lung damage in mice. 107 (2), 161. Liu, Y., Zhang, T., Fang, H.H., 2005. Microbial community analysis and performance of a phosphate-removing activated sludge. Bioresour. Technol. 96 (11), 1205–1214. Lu, S., Ryu, S.H., Chung, B.S., Chung, Y.R., Park, W., Jeon, C.O., 2007. Simplicispira limi sp. nov., isolated from activated sludge. Int. J. Syst. Evol. Microbiol. 57 (Pt 1), 31–34. Lu, Y., Yuan, T., Yun, S.H., Wang, W., Wu, Q., Kannan, K., 2010. Occurrence of cyclic and linear siloxanes in indoor dust from China, and implications for human exposures. Environ. Sci. Technol. 44 (16), 6081–6087. McGoldrick, D.J., Chan, C., Drouillard, K.G., Keir, M.J., Clark, M.G., Backus, S.M., 2014a. Concentrations and trophic magnification of cyclic siloxanes in aquatic biota from the Western Basin of Lake Erie, Canada. Environ. Pollut. 186, 141–148. McGoldrick, D.J., Chan, C., Drouillard, K.G., Keir, M.J., Clark, M.G., Backus, S.M.J.E.p., 2014b. Concentrations and trophic magnification of cyclic siloxanes in aquatic biota from the Western Basin of Lake Erie, Canada. 186, 141–148. Parker, W.J., Shi, J., Fendinger, N.J., Monteith, H.D., Chandra, G.J.E.T., 1999. Pilot plant study to assess the fate of two volatile methyl siloxane compounds during municipal wastewater treatment. Chemistry 18 (2), 172–181. Pieri, F., Katsoyiannis, A., Martellini, T., Hughes, D., Jones, K.C., Cincinelli, A., 2013. Occurrence of linear and cyclic volatile methyl siloxanes in indoor air samples (UK and Italy) and their isotopic characterization. Environ. Int. 59, 363–371. Quast, C., Pruesse, E., Yilmaz, P., Gerken, J., Schweer, T., Yarza, P., Peplies, J., Glockner, F.O., 2013. The SILVA ribosomal RNA gene database project: improved data processing and web-based tools. Nucleic Acids Res. 41 (Database issue), D590–D596. Sanchís, J., Martínez, E., Ginebreda, A., Farré, M., Barceló, D.J.S.o.t.T.E., 2013. Occurrence of linear and cyclic volatile methylsiloxanes in wastewater, surface water and sediments from Catalonia. 443, 530–538. Sawada, E., Satoh, T., Kitamura, H., 1978. Purification and properties of a dissimilatory nitrite reductase of a denitrifying phototrophic bacterium. Plant Cell Physiol. 19 (8), 1339–1351. Schloss, P.D., Westcott, S.L., Ryabin, T., Hall, J.R., Hartmann, M., Hollister, E.B., Lesniewski, R.A., Oakley, B.B., Parks, D.H., Robinson, C.J.J.A., microbiology, e., 2009. Introducing mothur: open-source, platform-independent, community-supported software for describing and comparing microbial communities. 75 (23),

7537–7541. Steiner, A., McLaren, D., Forster, C.J.W.R., 1976. The Nature of Activated Sludge Flocs. vol. 10(1). pp. 25–30. Stevens, C., Powell, D.E., Mäkelä, P., Karman, C.J.M.p.b., 2001. Fate and effects of polydimethylsiloxane (PDMS) in marine environments. 42 (7), 536–543. Stivala, S.S., 1980. Structure Vs Stability in Polymer Degradation. Polym. Eng. Sci. 20 (10), 654–661. Wagner, M., Erhart, R., Manz, W., Amann, R., Lemmer, H., Wedi, D., Schleifer, K.H., 1994. Development of an rRNA-targeted oligonucleotide probe specific for the genus Acinetobacter and its application for in situ monitoring in activated sludge. Appl. Environ. Microbiol. 60 (3), 792–800. Wang, D.G., Norwood, W., Alaee, M., Byer, J.D., Brimble, S., 2013a. Review of recent advances in research on the toxicity, detection, occurrence and fate of cyclic volatile methyl siloxanes in the environment. Chemosphere 93 (5), 711–725. Wang, D.G., Steer, H., Tait, T., Williams, Z., Pacepavicius, G., Young, T., Ng, T., Smyth, S.A., Kinsman, L., Alaee, M., 2013b. Concentrations of cyclic volatile methylsiloxanes in biosolid amended soil, influent, effluent, receiving water, and sediment of wastewater treatment plants in Canada. Chemosphere 93 (5), 766–773. Wang, D.-G., Aggarwal, M., Tait, T., Brimble, S., Pacepavicius, G., Kinsman, L., Theocharides, M., Smyth, S.A., Alaee, M.J.W.r., 2015. Fate of anthropogenic cyclic volatile methylsiloxanes in a wastewater treatment plant. 72, 209–217. Waring, M.S., Wells, J.R., 2015. Volatile organic compound conversion by ozone, hydroxyl radicals, and nitrate radicals in residential indoor air: magnitudes and impacts of oxidant sources. Atmos. Environ. 106, 382–391. Widyastuti, Y., Febrisiantosa, A.J.F., Sciences, N., 2014. The role of lactic acid bacteria in milk. fermentation 5 (04), 435. Wilen, B.M., Lumley, D., Mattsson, A., Mino, T., 2008. Relationship between floc composition and flocculation and settling properties studied at a full scale activated sludge plant. Water Res. 42 (16), 4404–4418. Xu, L., Shi, Y.L., Wang, T., Dong, Z.R., Su, W.P., Cai, Y.Q., 2012. Methyl Siloxanes in Environmental Matrices around a Siloxane Production Facility, and their distribution and Elimination in Plasma of Exposed Population. Environ. Sci. Technol. 46 (21), 11718–11726. Xu, S., Chandra, G.J.E.s., 1999. Fate of cyclic methylsiloxanes in soils. 2. Rates of degradation and volatilization. Technology 33 (22), 4034–4039. Xu, S., Lehmann, R.G., Miller, J.R., Chandra, G.J.E.s., technology, 1998. Degradation of Polydimethylsiloxanes (Silicones) as Influenced by Clay Minerals. vol. 32(9). pp. 1199–1206. Yamashita, T., Yamamoto-Ikemoto, R., 2014. Nitrogen and phosphorus removal from wastewater treatment plant effluent via bacterial sulfate reduction in an anoxic bioreactor packed with wood and iron. Int. J. Environ. Res. Public Health 11 (9), 9835–9853. Zhang, Z., Qi, H., Ren, N., Li, Y., Gao, D., Kannan, K., 2011. Survey of cyclic and linear siloxanes in sediment from the Songhua River and in sewage sludge from wastewater treatment plants, Northeastern China. Arch. Environ. Contam. Toxicol. 60 (2), 204–211. Zheng, C., Zhao, L., Zhou, X., Fu, Z., Li, A., 2013. Treatment technologies for organic wastewater. In: Water Treatment. Intech.

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