Cultivation of a versatile manganese-oxidizing aerobic granular sludge for removal of organic micropollutants from wastewater

Science of the Total Environment 690 (2019) 417–425

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Cultivation of a versatile manganese-oxidizing aerobic granular sludge for removal of organic micropollutants from wastewater Zhanfei He a, Qingying Zhang a, Zhen Wei a, Yuanhai Zhao a, Xiangliang Pan a,b,⁎ a b

Key Laboratory of Microbial Technology for Industrial Pollution Control of Zhejiang Province, College of Environment, Zhejiang University of Technology, Hangzhou, China Xinjiang Key Laboratory of Environmental Pollution and Bioremediation, Xinjiang Institute of Ecology and Geography, Chinese Academy of Sciences, Urumqi, China

H I G H L I G H T S • Manganese-oxidizing aerobic granular sludge (Mn-AGS) was successfully cultivated. • The degradation rates of most organic micropollutants were improved by Mn-AGS. • Dichlorophenyl phosphine (DCPP) was degraded faster by AGS than Mn-AGS. • Microbial activity of DCPP degradation might be inhibited by manganese oxides.

a r t i c l e

i n f o

Article history: Received 11 May 2019 Received in revised form 28 June 2019 Accepted 29 June 2019 Available online 02 July 2019 Editor: Damia Barcelo Keywords: Organic micropollutants (OMPs) Biodegradation Aerobic granular sludge (AGS) Biogenic manganese oxides (bio-MnOx) Sludge granulation

G R A P H I C A L

A B S T R A C T

manganese-oxidizing aerobic granular sludge DCPP BPA

Microbes EE2

CAP

organic micropollutants

inhibition?

CO2

TC IM

Mn(III,IV) r edox

BPA: Bisphenol A EE2:17α-Ethinylestradiol TC: Tetracycline CAP: Chloramphenicol IM: Imazethapyr DCPP: Dichlorophenyl phosphine

Mn(II) O2

a b s t r a c t Organic micropollutants (OMPs) are frequently detected in water and wastewater, and have attracted wide attention due to potential adverse effects on ecosystems and human health. In this work, manganese-oxidizing aerobic granular sludge (Mn-AGS) was successfully cultivated and applied to remove OMPs from wastewater. Biogenic manganese (III,IV) oxides (bio-MnOx) were generated and accumulated to 22.0–28.3 mg Mn/g SS in the final sludge. Neither the addition of allochthonous manganese-oxidizing bacteria (MnOB; Pseudomonas putida MnB1) nor the reduction in hydraulic retention time (HRT) facilitated the cultivation of Mn-AGS. Batch experiments of OMPs degradation indicated that Mn-AGS significantly improved (1.3–3.9 times) degradation rates of most OMPs. Removal rates of bisphenol A (BPA), 17α ethinylestradiol (EE2), tetracycline (TC), and chloramphenicol (CAP) were 3.0–12.6 μg/h/g SS by the traditional AGS and 8.0–16.3 μg/h/g SS by Mn-AGS; those of imazethapyr (IM) were relatively high, 64.7 ± 0.1 and 127.8 ± 2.5 μg/h/g SS by AGS and Mn-AGS, respectively. However, degradation of dichlorophenyl phosphine (DCPP) was slower by Mn-AGS than AGS, 9.0 ± 0.4 vs. 21.2 ± 0.9 μg/h/g SS, possibly due to inhibition of microbial activity by bio-MnOx. This work provides a promising method for treating OMPs in organic wastewater, but the possible inhibition of microbes by bio-MnOx should be noted. © 2019 Elsevier B.V. All rights reserved.

1. Introduction

⁎ Corresponding author at: College of Environment, Zhejiang University of Technology, Hangzhou 310014, China. E-mail address: [email protected] (X. Pan).

https://doi.org/10.1016/j.scitotenv.2019.06.509 0048-9697/© 2019 Elsevier B.V. All rights reserved.

Organic micropollutants (OMPs), such as hormones, antibiotics, pharmaceuticals, and personal care products, are frequently found in water and wastewater, and have gained increasing attention from both the scientific community as well as the general public (Verlicchi

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et al., 2012; Mahamallik et al., 2015; Bilal et al., 2019; Moreira et al., 2019; Tran et al., 2019). OMPs are introduced to natural water by many possible routes, including discharge of treated effluents from wastewater treatment plants, sewer leakage/overflow, urban and agricultural runoff, and illicit discharge (Le et al., 2018; Tran et al., 2019). OMPs may adversely affect aquatic ecosystems and human health and they can bioaccumulate through the food chain (Rozas et al., 2016; Guo et al., 2017; Tran et al., 2018a). Moreover, some OMPs, such as bisphenol A (BPA), are known endocrine-disrupting chemicals (EDCs) which animal studies have shown to cause cancer and negatively affect the prostate, kidney, breast, and reproductive and immune systems (Lane et al., 2015; Bilal et al., 2017; Barrios-Estrada et al., 2018). Biogenic manganese oxide (bio-MnOx), a product of manganese (Mn) oxidization by manganese-oxidizing bacteria (MnOB) or fungi, is well known for its high redox potential of 500–600 mV and high adsorption capacity due to large specific surface area (Furgal et al., 2015). Mn(III,IV) in bio-MnOx can oxidize a variety of OMPs, such as BPA, tetracycline (TC) and diclofenac (Mahamallik et al., 2015; Wang et al., 2017), and is reduced to Mn(II) which can be re-oxidized into bio-MnOx by MnOB (Tran et al., 2018b). Many OMPs may be simultaneously degraded by MnOB or manganese-oxidizing fungi and their product, bio-MnOx (Auriol et al., 2008; Tran et al., 2010; Tran et al., 2013). Previous studies have shown a strong correlation between 17α ethinylestradiol (EE2) degradation and bio-MnOx generation in MnOB- and Mn(II)-amended systems, implying that EE2 degradation is mainly dependent on bio-MnOx (Sabirova et al., 2008). Very recently, experimental results indicated that presence of MnOB, Pseudomonas putida MnB1 (P. putida), greatly (15-fold) improved the EE2 degradation by bio-MnOx, suggesting that MnOB might also be involved in the chemical reaction of bio-MnOx (Tran et al., 2018b). Aerobic granular sludge (AGS) is a high-efficiency biotechnology for wastewater treatment because of its excellent settleability, high biological activity, and high tolerance for toxic compounds. Organic pollutants, including some OMPs, can be effectively removed by AGS technology through microbial degradation or surface adsorption. Wang et al. (2019) detected five antibiotics in a piggery wastewater, including tetracycline, kanamycin, ciprofloxacin, ampicillin, and erythromycin, and found that these antibiotics could be removed by AGS with a total rate of 88.4 ± 4.5%. AGS was also used to treat some pesticides in wastewater, such as 2,4 dichlorophenoxyacetic acid (Ma et al., 2012). Effluents from wastewater treatment plants are major sources of OMPs in natural environments, as many OMPs are difficult to efficiently remove through the conventional activated sludge process (Rozas et al., 2016; Tran and Gin, 2017; Le et al., 2018; Rasheed et al., 2019; Tran et al., 2019). Although membrane filtration and advanced oxidation processes can remove OMPs, the cost of implementation is prohibitive, especially for agriculture and animal husbandry operations (Furgal et al., 2015; Tolboom et al., 2019). Manganese-oxidizing aerobic granular sludge (Mn-AGS) is a promising new method for degrading and removing OMPs from wastewater that combines the advantages of bio-MnOx and AGS. Compared to bio-MnOx generation by the pure culture method (Furgal et al., 2015), Mn-AGS is more promising for practical engineering applications because of its excellent settleability and high tolerance for toxic compounds. Cultivating of Mn-AGS will be the most important step for this new technology, since large amounts of sludge is required for both scientific research and practical applications. Reactor configuration and hydraulic conditions affecting AGS (Adav et al., 2008), may also influence Mn-AGS cultivation, and should thus be considered during Mn-AGS granulation and reactor operation. In the present study, Mn-AGS was obtained by continuously adding Mn(II) to three sequencing batch reactors (SBRs), and AGS was cultivated under the same conditions without Mn(II). The effects of the MnOB addition and hydraulic retention time (HRT) reduction on MnAGS cultivation were investigated, and extracellular polymeric substances (EPS) analysis, X-ray photoelectron spectroscopy (XPS) analysis, and microbial analysis were performed on both the inoculum and

the final sludge. The Mn-AGS and AGS obtained through the above methods were investigated for usefulness in OMPs degradation, and possible mechanisms were introduced according to the findings. 2. Materials and methods 2.1. Seed sludge and medium The seed sludge (activated sludge floc) was collected from a secondary sedimentation tank of a WWTP in Hangzhou City, China. The activated sludge was settled for 1.0 h and 100 mL of concentrated sludge was seeded in each reactor. Two media were involved in this work: acetate medium and glucose medium. The acetate medium contained 1300–1600 mg/L NaAc, 229.3 mg/L NH4Cl, 43.9 mg/L KH2PO4, 0.29 g/L CaCl2·2H2O, 0.82 g/L MgSO4·7H2O, 0 or 22.9 mg/L MnCl2 (0 mg/L for AGS and 22.9 mg/L for Mn-AGS), and 0.3 mL/L trace element solution I (see Table S1). The glucose medium consisted of 1.0 g/L glucose, 0.5 g/L yeast extract, 0.5 g/L casamino acids, 0.29 g/L CaCl2·2H2O, 0.82 g/L MgSO4·7H2O, 22.9 mg/L MnCl2, 1.0 mg/L FeCl3, and 1.0 mL/L trace element solution II (see Table S1). pH of the initial influent was adjusted at 7.0–7.2. 2.2. Reactor setup and operation In order to cultivate granular sludge of AGS and Mn-AGS, four identical SBRs were run continuously for 55 days. The four SBRs operated under four different conditions and were labeled “AGS”, “Mn-AGS”, “Mn-AGS(HRT)”, and “Mn-AGS(MnOB)”. Detailed experimental settings of the four reactors are provided in Table 1. Briefly, in the AGS reactor Mn(II) was not supplied and the reactor was used to cultivate AGS as a control group; in the Mn-AGS reactor 10 mg/L of Mn(II) was continuously added to influent to obtain Mn-AGS; in the Mn-AGS(HRT) reactor hydraulic retention time (HRT) was reduced from 18 h to 9 h; and in the Mn-AGS(MnOB) reactor 100 mL cell suspension of a typical manganese-oxidizing bacterium, Pseudomonas putida (P. putida) MnB1, was added daily. The operation period of the Mn-AGS(MnOB) reactor was divided into two phases: in Phase I (days 1–22) glucose medium was used; and in Phase II (days 23–55), acetate medium was used. Cell suspension of P. putida was prepared in a liquid culture using the glucose medium above (He et al., 2019c). Fig. 1 shows a schematic diagram of the SBRs. Each reactor had a working volume of 1.8 L, and the volume exchange ratio was 2/3 in the Mn-AGS(HRT) reactor (HRT = 9 h) and 1/3 in the other reactors (HRT = 18 h). One SBR cycle was 6 h, including 10 min of medium feeding, 5.5 h of aeration, 10 to 1 min of settling (10 min on days 1–10, 5 min on days 11–30, and 1 min on days 31–55), 5 min of effluent discharge, and 5 to 14 min of idle. An air pump conveyed air into the reactor bottom at a rate of 200 L/h, and an air-conditioner was employed to keep the temperature near 25 °C. Chemical oxygen demand (COD), pH, and Mn(II) in both influent and effluent were sampled and measured every day. 2.3. Batch experiments To assess the degradation potential of OMPs by Mn-AGS and AGS, six pollutants were tested, including two endocrine-disrupting chemicals (BPA and EE2), two antibacterial agents (TC and chloramphenicol (CAP)), and two herbicides (imazethapyr (IM) and dichlorophenyl phosphine (DCPP)). These OMPs were widely applied and relatively difficult to be biodegrade. AGS was used as a control group and all tests were performed in triplicate. Batch experiments were carried out in 150-mL conical flasks. First, Mn-AGS or AGS sludge was harvested from the above reactors on day 55 and rinsed with fresh acetate medium. Then, each flask was loaded with 10 mL sludge and 40 mL fresh acetate medium with pollutants. The initial pollutant (BPA, EE2, TC, CAP, IM, and DCPP) concentrations were 2.0, 2.0, 1.0, 1.0, 2.0, and

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Table 1 Experimental settings of the four reactors in this work. Reactora

Mn(II) (mg/L)

HRT (h)

MnOB addition

0 10 10 10

18 18 9 18

Not added Not added Not added Added

AGS Mn-AGS Mn-AGS(HRT) Mn-AGS(MnOB) a

Medium Acetate medium Acetate medium Acetate medium Glucose medium (days 1–22); acetate medium (days 23–55)

50-mL influent and effluent were sampled from each reactor every day, and 20-mL sludge was collected regularly.

2.0 mg/L, respectively. The flasks were then placed on a shaking table at 150 rpm and 25 ± 0.5 °C. 2 mL suspension samples were taken and the supernatant was filtered (0.22 μm) every 2 h to measure concentrations of the corresponding pollutants. Sludge was collected and dried at 105 °C to measure suspended solids (SS) after experiments. Bio-MnOx was prepared with P. putida MnB1 and purified with deionized water as previously described (Tran et al., 2018b). 2.4. EPS extraction and analysis Sludge samples were taken before and after reactor operation to extract and analyze EPS. The EPS was extracted using the formaldehydeNaOH method as previously described (Adav et al., 2008). The contents of protein (PN) and polysaccharide (PS) in EPS extracts were measured according to the Lowry method (Lowry et al., 1951) and the phenol sulfuric acid method (Masuko et al., 2005), respectively. The EPS content was calculated by adding PN and PS contents. The spectra of three-dimensional excitation-emission matrix (3D-EEM) fluorescence of the extracted EPS were recorded with a spectrofluorophotometer (RF-6000, Shimadzu, Japan). The ranges of the excitation and emission wavelengths were set at 200–550 nm and 200–600 nm, respectively, and both scanning intervals were 2.0 nm. The excitation and emission slits were set at 5.0 and 10.0 nm, respectively. A spectrum of deionized water was measured as a blank. The EEM data were presented using Matlab 2014a (The Mathworks Inc., USA) and a logarithmic plot was adopted to show the fluorescence intensity.

Air pump Flowmeter

Bioreactor

Peristaltic pump Outlet

Gauze Bubble Diffuser

Influent tank

Inlet

Effluent tank

Fig. 1. Schematic diagram of the reactors used in this work.

2.5. Analytical methods COD, SS, and volatile suspended solids (VSS) were determined according to the standard methods for examination of water and wastewater (APHA, 2005). Liquid pH was measured with a FE20 pH meter (Mettler-Toledo, Shanghai). Mn(II) was monitored using the potassium periodate spectrophotometric method, and Mn(III,IV) was determined by the LBB quantitative method as described in our previous work (He et al., 2019c). Organic pollutants of BPA, EE2, TC, CAP, IM, and DCPP were measured using a high performance liquid chromatograph (HPLC) (LC-20A, Shimadzu, Japan) equipped with a model SPD-20A UV detector and a Shimadzu InertSustain C18 column (5 μm, 4.6 × 150 mm). The mobile phase consisted of acetonitrile/water (40:60) (BPA), acetonitrile/water (70:30) (EE2), acetonitrile/water (35:65) and 0.01 M KH2PO4 (TC and CAP), acetonitrile/water/acetic acid (40:59.8:0.2) (IM), or acetonitrile/water (40:60) (DCPP), and the flow rate was set at 1.0 mL/min. The column temperature was 40 °C (TC, CAP, and IM) or 30 °C (BPA, EE2 and DCPP), and the detection wavelength was 280 nm (BPA and EE2), 375 nm (TC), 277 nm (CAP), 254 nm (IM) or 250 nm (DCPP). Limits of detection (LOD) and quantitation (LOQ) for BPA, EE2, TC, CAP, IM, and DCPP were 0.001 and 0.003, 0.013 and 0.039, 0.009 and 0.027, 0.002 and 0.005, 0.003 and 0.008, and 0.004 and 0.013 mg/L, respectively, estimated from the standard deviation and the calibration curve (Saadati et al., 2013). XPS characterization was performed to analyze the valence state of Mn in different sludges. The sludge samples were vacuum-dried and ground, and their XPS spectra were recorded using an X-ray photoelectron spectrometer (ESCALAB 250Xi, ThermoFischer, USA) (He et al., 2019c). The spectra of C 1s and Mn 2p were scanned, and the charge effect on Mn 2p was adjusted by C1s (284.80 eV). The XPS data of Mn 2p were fitted with XPS Peak4.1 software, and the Shirley function was used to subtract the background (Kim et al., 2014). For microbial community analysis, sludge was sampled before and after experiments (on days 1 and 55). Due to a medium change in the Mn-AGS(MnOB) reactor, one more sludge sample was added on day 22; the samples from days 22 and 55 were marked as Mn-AGS (MnOB)#1 and Mn-AGS(MnOB)#2, respectively. DNA extraction was performed using the Power Soil DNA isolation kit (MoBio, United States) in accordance with the manufacturer's instructions. The bacterial 16S rRNA genes (V3–V4 regions) of the template DNA were amplified by 338F/806R primers tagged with barcode sequences as previously described (Shen et al., 2017; He et al., 2019b). The PCR procedure included the following steps: initial denaturation (94 °C, 3 min), 25 cycles of dissociation (94 °C, 30 s), annealing (56 °C, 30 s) and extension (72 °C, 30 s), followed by a final extension (72 °C, 5 min). The PCR products were purified and paired-end sequenced on an Illumina platform. The obtained sequences were analyzed by Shanghai Majorbio Biotech Co., Ltd. using a standard procedure of Quantitative Insights Into Microbial Ecology (QIIME) (Caporaso et al., 2010). The raw paired reads were merged using FLASH software (Magoč and Salzberg, 2011). Low-quality sequences (b200 bp, containing ambiguous bases, quality score b 20) were discarded by the software QIIME (Caporaso et al., 2010). Chimeras were checked and removed by the USEARCH program (Edgar et al., 2011). The remaining high-quality sequences were then clustered into operational taxonomic units (OTUs) with a similarity

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threshold of 97%. Singletons, the unique OTUs that only appeared once, were removed and the other non-singleton OTUs were assigned against the SILVA database (release 132). Clustering and principal coordinate analyses (PCoA) were conducted using the I-sanger platform (https:// www.i-sanger.com) to reveal correlations of the microbial community among different samples. 3. Results and discussion 3.1. Reactor performance

1250

During the operating period, COD, pH, and Mn(II) were monitored daily in influent and effluent. As shown in Fig. 2a, COD was efficiently removed in all the reactors after several days' adjustment. Effluent COD was stable at 50 ± 11 mg/L after day 10, and removal percentage was approximately 96% (Fig. S1). Comparisons of results between reactors AGS and Mn-AGS showed that the addition of 10 mg/L Mn(II) did not cause any significant impact on COD removal. In Fig. 2b, pH increased from influent 7.1 ± 0.1 to effluent 8.9 ± 0.2 with the acetate medium and to effluent 7.7 ± 0.2 with the glucose medium. The pH elevation in acetate medium can be attributed to oxidation of acetate (Eq. (1)), which produced bicarbonate (HCO− 3 ) and increased the solution pH. Glucose oxidation (Eq. (2)) does not generate a base, and the small increases in pH may have been caused by the oxidation of casamino acids or yeast extract in the medium. ð1Þ

1000

Influent (NaAc) Influent (Glucose) Eff. AGS Eff. Mn-AGS Eff. Mn-AGS (HRT) Eff. Mn-AGS (MnOB)

750 500 250 0 0

10

20

30

60

(b) 9.0 8.5

Influent (NaAc) Influent (Glucose) Eff. AGS Eff. Mn-AGS Eff. Mn-AGS (HRT) Eff. Mn-AGS (MnOB)

8.0

ð2Þ

7.0 0

10

20

30

40

50

60

Time (d) 14

(c) 12 10 Mn(II) (mg Mn/L)

Fig. 2c indicated that effluent Mn(II) was fluctuant at first and remained stable after day 30. The microbial acclimatization and shortening of the SBR settling time (from 10 to 1 min) during the first 30 days may cause fluctuation in effluent Mn(II). Effluent Mn(II) was 2.9 ± 0.3 mg/L in reactor Mn-AGS(HRT) and 1.7 ± 0.2 mg/L in reactors Mn-AGS and Mn-AGS(MnOB) after day 30, and their removal percentages were 71.2 ± 2.5% and 82.9 ± 2.4%, respectively. HRT in reactor Mn-AGS(HRT) was half that of the other reactors (9 h vs. 18 h), meaning that Mn(II) loading was double in reactor Mn-AGS(HRT), resulting in low removal percentages of Mn(II). Considering the large amount of bicarbonate in the reactors, influent Mn(II) would be precipitated with bicarbonate to form manganese carbonate (rhodochrosite, MnCO3). Rhodochrosite can also be oxidized into bio-MnOx by microbes (Wang et al., 2015a). The inoculum and the final sludge (day 55) were stained with LBB, and their photographs are shown in Fig. S2. The sludge samples from the reactors with added Mn(II) were stained blue by LBB, while the inoculum and AGS samples were not stained. Further, Mn(III,IV) content in the final sludges was assessed using the LBB quantitative method (Li et al., 2014; He et al., 2019c), and the values were 0.57 ± 0.02 (AGS), 28.3 ± 0.9 (Mn-AGS), 22.0 ± 0.9 (Mn-AGS(HRT)), and 23.3 ± 1.7 mg Mn/g SS (Mn-AGS(MnOB)). It clearly showed that bioaugmentation of adding cell suspension (P. putida MnB1) into the reactor did not increase bio-MnOx generation. Moreover, the granule shapes in all the final sludges were clear (Fig. S2), and ratios of SVI30/SVI5 were very close to 1.0 (Table S2), indicating that sludge granulation had been completed (Pronk et al., 2015).

50

9.5

7.5 C6 H12 O6 þ 6O2 →6CO2 þ 6H2 O

40

Time (d)

pH

CH3 COO– þ 2O2 →HCO3 – þ CO2 þ H2 O

(a)

1500

COD (mg COD/L)

420

Influent (NaAc) Influent (Glucose) Eff. Mn-AGS Eff. Mn-AGS (HRT) Eff. Mn-AGS (MnOB)

8 6 4 2 0 0

10

20

30

40

50

60

Time (d) Fig. 2. Influent and effluent COD (a), pH (b), and Mn(II) (c) in different reactors. In days 1–22, the glucose medium was used in reactor Mn-AGS(MnOB) and the other reactors shared one mother acetate medium; in days 23–55, all of the reactors shared one mother acetate medium.

3.2. EPS content and EEM spectra Aerobic granulation was greatly promoted by EPS (Adav et al., 2008), and EPS of the sludge was extracted and analyzed. As shown in Fig. 3a, PN content was much higher (6–9 times) than PS in all sludges, and it was known to support granulation (Zhang et al., 2019). EPS contents,

both PN and PS, increased during granulation in each reactor, and were highest in the reactor with a short HRT, consistent with previous results (Wang et al., 2015b). The EPS contents in reactors AGS, Mn-

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421

Fig. 3. Components (a) and EEM fluorescence spectra of EPS extracted from inoculum (b) and the final sludge (reactors AGS (c), Mn-AGS (d), Mn-AGS(HRT) (e), and Mn-AGS(MnOB) (f)).

AGS, and Mn-AGS(MnOB) were similar, indicating that addition of Mn (II) or MnOB did not significantly change EPS components or content. The EEM fluorescence spectra of EPS are shown in Figs. 3b–f. Four peaks were recognized in the samples and labeled as A, B, C, and D. Peaks A and B were adjacent to each other and located at excitation/

emission (Ex./Em.) of 250–310/320–470 nm and 310–350/340– 470 nm, respectively. Both of them could be assigned to tryptophan and protein-like substances, according to the fluorescence regional integration reported by Chen et al. (2003). Peak C was located at Ex./Em. of 220–250/320–450 nm, corresponding to aromatic protein-like

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amended reactors, and they may be directly oxidized from dissolved Mn(II) or indirectly oxidized from solid Mn(II) (rhodochrosite). Furthermore, the results showed that the addition of MnOB did not promote generation of bio-MnOx.

substances. Peak D was observed at Ex./Em. of 350–450/430–570 nm, which was related to humic acid-like substances (Chen et al., 2003). On the basis of the fluorescence intensity, the EEM results indicated that large numbers of proteins and few humic acid-like substances were present in the EPS. Comparisons of the EEM spectra revealed that humic acid-like substances distinctly decreased after granulation, especially in the reactors amended with Mn(II). Sludge granulation and bio-MnOx generation may reduce generation of humic acid-like substances or stimulate their consumption.

3.4. Microbial community characteristics The granulation and biological processes, including COD degradation and Mn(II) oxidation, were determined by the sludge microbial community which was analyzed based on the 16S rRNA gene sequences. As shown in Fig. S3, inoculum was jointly dominated by phyla of Chloroflexi (36.8%), Proteobacteria (22.5%), and Actinobacteria (19.3%), whereas the final sludge was dominated by Proteobacteria (72.2%–85.5%) and Bacteroidetes (7.1%–31.2%). The relative abundance of microbial communities at genus level is shown in Fig. 5. The genera in the inoculum was rich and no one genus accounted for N10%. Other sludge was dominated by several genera, and genus Paracoccus was abundant in all the cultivated sludge. Paracoccus members, known for their ability to use a wide range of organic matter including glucose and acetate (McGinnis et al., 2015), were abundant in both glucose (sample of Mn-AGS(MnOB)#1) and acetate media (the final sludge). Some genera known to oxidize Mn (II), such as Bacillus, Pseudomonas, Acinetobacter, Sphingobacterium, Pedomicrobium, and Roseobacter (Van Waasbergen et al., 1996; Ridge et al., 2007; Andeer et al., 2015; Li et al., 2016), were detected in the final sludge. They may have been responsible for the generation of bioMnOx in these sludges. Moreover, bacterial species of P. putida were

3.3. Mn valence state in granules The XPS spectra of the four final sludges were recorded, and the Mn 2p spectra are shown in Fig. 4. There was no discernible sign of Mn 2p in the AGS granules in this work (Fig. 4a), whereas Mn 2p spectra were apparent in the other three samples (Fig. 4b–d). These Mn 2p spectra were modeled as 2p1/2, 2p3/2, and shake-up satellite peaks as previously described (Ilton et al., 2016). Among them, Mn 2p3/2 spectra were fitted by three peaks located at 642.6–642.8 eV, 641.6–641.7 eV, and 640.8–640.9 eV, representing 2p3/2 spectra of Mn(IV), Mn(III), and Mn (II) (labeled in Fig. 4), respectively (He et al., 2019c). According to the peak area, the relative abundances of Mn(IV), Mn(III), and Mn(II) were 52.6%, 23.1%, and 24.3% in the Mn-AGS reactor, 56.4%, 20.1%, and 23.6% in the Mn-AGS(HRT) reactor and 47.9%, 20.9%, and 31.2% in the Mn-AGS(MnOB) reactor. These data further support the conclusion that bio-MnOx, Mn(III,IV) oxides, were produced in the Mn(II)-

(a)

(b) 2p3/2

Intensity

Intensity

2p1/2

AGS 660

Mn(IV) Mn(II)

Mn-AGS 655

650

645

640

635

660

655

Binding Energy (eV)

650

645

640

635

Binding Energy (eV)

(d)

(c) 2p3/2 2p1/2

2p1/2 Mn(III)

Intensity

Intensity

2p3/2

Mn(IV) shake-up satellite

Mn(II)

Mn-AGS(HRT) 660

Mn(III)

shake-up satellite

655

Mn(IV) shake-up satellite

Mn(III) Mn(II)

Mn-AGS(MnOB) 650

645

Binding Energy (eV)

640

635

660

655

650

645

Binding Energy (eV)

Fig. 4. XPS spectra of the final sludge in reactors AGS (a), Mn-AGS (b), Mn-AGS(HRT) (c), and Mn-AGS(MnOB) (d).

640

635

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Fig. 5. Clustering analysis and relative abundance of microbial community at genus level. Only top 50 genera were presented and other genera were lumped into “others”. (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

detected in the sludge sample from reactor Mn-AGG(MnOB), but only accounted for 0.06–0.09% of the microbial community. This indicates that the added P. putida MnB1 failed to thrive and grow in the reactor, even when glucose medium was provided in the beginning. It is therefore apparent that the performance of the MnOB-amended reactor (reactor Mn-AGS(MnOB)) was similar without the addition of MnOB (reactor Mn-AGS), since only a few P. putida MnB1 retained in the reactor. The results of clustering and PCoA analyses are given in Fig. 5 (the red lines in the right part) and Fig. S4, respectively. Both analyses showed that microbial community structure was similar in the final sludge from reactors AGS, Mn-AGS, and Mn-AGS(MnOB), indicating that additions of Mn(II) and MnOB did not cause large changes in the microbial community. HRT evidently influenced the microbial community; carbon sources greatly affected the microbial community, possibly because carbon metabolism is the most important primary metabolism of microorganisms (Luckner, 1984). 3.5. Potentials of OMPs degradation by Mn-AGS Three types of OMPs including two endocrine disrupting chemicals (BPA and EE2), two antibacterial agents (TC and CAP), and two herbicides (IM and DCPP), were tested for degradation by AGS and Mn-AGS, and the results are given in Fig. 6 and Fig. S5. As shown in Fig. 6, all tested OMPs could be removed by both AGS and Mn-AGS, and biosorption and biodegradation were responsible for the removal (Bilal et al., 2018; Rasheed et al., 2019). With the exception of IM (64.7 ± 0.1 and 127.8 ± 2.5 μg/ h/g SS in AGS and Mn-AGS, respectively), other pollutants including BPA, EE2, TC, and CAP, were degraded at low rates, 3.0–12.6 μg/h/g SS by AGS and 8.0–16.3 μg/h/g SS by Mn-AGS. Overall, the removal rates in the Mn-AGS group were higher (1.3–3.9 times) than those in AGS group. Improved OMPs degradation in Mn-AGS may be attributed to the highly active bio-MnOx, which has been reported to degrade BPA, EE2, and TC at high rates (Sabirova et al., 2008; Wang et al., 2015a; Wang et al., 2017). The kinetic constants of OMPs biodegradation by AGS were estimated to be 0.05 (BPA), 0.34 (EE2), 0.12 (TC), and 0.32 (CAP) L/

g SS/d, which were mostly lower than the kinetic constants of OMPs degradation by activated sludge (0.24–16.56 (BPA), 0.02–20 (EE2), and 0.44 L/g SS/d (TC)) (Tran et al., 2018a). Low mass transfer rate of OMPs in granules may slow the degrading rate by AGS. The kinetic constants of Mn-AGS were 0.13 (BPA), 0.56 (EE2), 0.98 (TC), and 0.47 (CAP) L/ g SS/d, and the degradation of TC was faster than by activated sludge (0.44 L/g SS/d). The results of DCPP were the opposite, however, as AGS showed a higher performance than Mn-AGS. Two supplementary experiments were subsequently carried out for groups AGS + bio-MnOx and bioMnOx to investigate the poor performance of Mn-AGS in DCPP degradation. An identical quantity of bio-MnOx (~1.0 g) was added to each group, and removal rates are shown in Fig. 6. The specific rates for the two groups were estimated by dividing removal rates (Fig. S5) by the AGS weight of the AGS + bio-MnOx group. As given in Fig. 6, the specific removal rates of DCPP were 21.2 ± 0.9, 9.0 ± 0.4, 35.7 ± 0.2, and 30.4 ± 0.4 μg/h/g SS in groups AGS, Mn-AGS, AGS + bio-MnOx, and bio-MnOx, respectively. Assuming that chemical degradation of DCPP by bio-MnOx was not majorly influenced by AGS addition, the DCPP removal rate by AGS was calculated to be 5.3 ± 0.4 μg/h/g SS (by subtracting 30.4 ± 0.4 μg/h/g SS from 35.7 ± 0.2 μg/h/g SS) in the presence of bio-MnOx (group AGS + bio-MnOx). This was much lower than in absence of bio-MnOx in group AGS (5.3 vs. 21.2 μg/h/g SS), suggesting that microbial activity of DCPP degradation in sludge was inhibited by bio-MnOx. Similar inhibition of microbial activity by bio-MnOx has also been observed in previous literature (Matsushita et al., 2018). Although bioMnOx in Mn-AGS can directly remove DCPP (Fig. 6), it may greatly inhibit biological activity of DCPP-degrading bacteria and thus lower the overall efficiency. Bio-MnOx may generate reactive oxygen species (ROS) in sludge and damage cellular enzymes, membranes, and DNA, similar to microbial inhibition by other metal oxides (Stankic et al., 2016). However, it was different to explain why DCPP-degrading bacteria were seriously damaged while others were not. In summary, most OMPs tested in this work were slowly degraded by microbes and their degradation rates were improved by Mn-AGS.

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Fig. 6. Specific degradation rates of OMPs by AGS and Mn-AGS. Two additional tests (groups AGS + Bio-MnOx and Bio-MnOx) were carried out to explain the unusual results of DCPP. Abbreviation: BPA, bisphenol A; EE2, 17α ethinylestradiol; TC, tetracycline hydrochloride; CAP, chloramphenicol; IM, imazethapyr; DCPP, dichlorophenyl phosphine.

Microbial and bio-MnOx were the two possible pathways for OMPs degradation by Mn-AGS (Fig. S6). Intermediate exchange may have occurred between the two pathways, since many OMPs were degraded by sequential reactions and many intermediates were produced (Wang et al., 2015a; He et al., 2019a). The produced bio-MnOx may inhibit microbial activity of OMPs degradation such as the DCPP degradation in this work, which was a disadvantage of this biotechnology. 4. Conclusion Mn-AGS was successfully obtained in SBRs after 55 days' operation with addition of Mn(II), and the content of bio-MnOx in the final sludge was 22.0–28.3 mg Mn/g SS. Neither the addition of MnOB (P. putida MnB1) nor the reduction of HRT (from 18 h to 9 h) facilitated cultivation of Mn-AGS; other strategies should be proposed in future studies. Degradation rates of most OMPs, including BPA, EE2, TC, CAP, and IM, were improved (1.3–3.9 times) by Mn-AGS compared to AGS, with the exception of DCPP. Biodegradation by microbes and chemical oxidation by bio-MnOx were the main pathways of OMPs degradation by Mn-AGS. Bio-MnOx in Mn-AGS improved the degradation of most OMPs but largely inhibited microbial activity of DCPP-degrading microorganisms, which led to a relatively low DCPP degradation rate by Mn-AGS. In general, Mn-AGS showed a promising method for simultaneously removing COD and OMPs from organic wastewaters, but some OMPs, like DCPP, are not feasible. Acknowledgements This work was funded by the National Key Research and Development Program of China (2018YFC1802902) and the National Natural Science Foundation of China (41701274, U1503281 and U1703243).

Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2019.06.509.

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