Degradation of 17β-estradiol by Novosphingobium sp. ES2-1 in aqueous solution contaminated with tetracyclines

Degradation of 17β-estradiol by Novosphingobium sp. ES2-1 in aqueous solution contaminated with tetracyclines

Journal Pre-proof Degradation of 17β-estradiol by Novosphingobium sp. ES2-1 in aqueous solution contaminated with tetracyclines Shunyao Li, Juan Liu, ...

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Journal Pre-proof Degradation of 17β-estradiol by Novosphingobium sp. ES2-1 in aqueous solution contaminated with tetracyclines Shunyao Li, Juan Liu, Kai Sun, Zhiyao Yang, Wanting Ling PII:

S0269-7491(19)35839-7

DOI:

https://doi.org/10.1016/j.envpol.2020.114063

Reference:

ENPO 114063

To appear in:

Environmental Pollution

Received Date: 8 October 2019 Revised Date:

15 January 2020

Accepted Date: 22 January 2020

Please cite this article as: Li, S., Liu, J., Sun, K., Yang, Z., Ling, W., Degradation of 17β-estradiol by Novosphingobium sp. ES2-1 in aqueous solution contaminated with tetracyclines, Environmental Pollution (2020), doi: https://doi.org/10.1016/j.envpol.2020.114063. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2020 Published by Elsevier Ltd.

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Graphical abstract

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Degradation of 17β-estradiol by Novosphingobium sp. ES2-1 in

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aqueous solution contaminated with tetracyclines

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Shunyao Lia,1, Juan Liua,1, Kai Sun2, Zhiyao Yang1, Wanting Ling*,1

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and Environmental Sciences, Nanjing Agricultural University, Nanjing 210095, China

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Prevention, School of Resources and Environment, Anhui Agricultural University, Hefei,

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Institute of Organic Contaminant Control and Soil Remediation, College of Resources

Anhui Province Key Laboratory of Farmland Ecological Conservation and Pollution

230036, China

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*Corresponding author: Wanting Ling, Dr/Professor

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Address: Weigang Road 1, Nanjing 210095, China

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Tel: +86-25-84395194

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Orcid: 0000-0002-2376-7760

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E-mail: [email protected]

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a

These authors contributed equally to this paper.

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Abstract

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17β-estradiol (E2) often coexists with tetracyclines (TCs) in wastewater lagoons at

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intensive breeding farms, threatening the quality of surrounding water bodies. Microbial

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degradation is vital in E2 removal, but it is unclear how TCs affect E2 biodegradation.

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This primary study investigated the mechanisms of E2 degradation by Novosphingobium

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sp. ES2-1 in the presence of TCs and assessed the removal efficiency of E2 by strain

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ES2-1 in natural waters containing TCs. E2 biodegradation was unaffected at TCs

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concentrations below 0.1 mg L−1 yet significantly inhibited at TCs above 10 mg L−1. As

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elevation of TCs, E2 biodegradation rate constant decreased, and the biodegradation

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kinetics equation gradually deviated from the pseudo-first-order dynamics model.

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Importantly, the presence of TCs, especially at high-level concentrations, significantly

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hindered E2 ring-opening process but promoted the condensation of some phenolic

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ring-opening products with NH3, thereby increasing the abundance of pyridine derivatives,

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which were difficult to decompose over time. Additionally, strain ES2-1 could remove

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52.1–100% of nature estrogens in TCs-contaminated natural waters within 7 d. Results

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revealed the mechanisms of TCs in E2 biodegradation and the performance of a functional

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strain in estrogen removal in realistic TCs-contaminated aqueous solution.

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Keywords: Antibiotics; Estrogens; Degradation; Sphingomonad; Combined pollution

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Capsule: TCs may inhibit the E2 degradation by strain ES2-1, and disturb the cleavage of

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E2, reducing the abundance of ring-opening products.

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1. Introduction

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17β-estradiol (E2) is an estrogen of priority concern due to its high endocrine activity

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(Bradley et al., 2009). If allowed to accumulate in the ecosystem and enter the human food

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chain, E2 can not only disturb the sex ratio balance of aquatic wildlife but induce human

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cancer and increase the probability of viral infections (Wocławek-Potocka et al., 2013;

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Arnold et al., 2014; Tetreault et al., 2011; Rose et al., 2013; Yang et al., 2018). Both the

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US Environmental Protection Agency and the European Commission have included

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estrogens in regulatory lists due to their potential side effects (US-EPA, 2016; EE

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Commission, 2013). Livestock waste is a major source of estrogens. It was estimated that

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estrogens discharged by livestock in the United States and European Union reached

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83,000 kg per year (Adeel et al., 2017). The UK average annual emission of E2 in 2013

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reached 570 kg (Ray et al., 2013).

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Storage of liquid manure, manure spreading, and surface runoff provide convenient

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modes for E2 in livestock waste to enter water systems, leading to high detection

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frequency of E2 in surface waters, with levels ranging from ng L−1 to µg L−1 (Beck et al.,

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2006; Kjær et al., 2007; Pal et al., 2010; Rocha et al., 2012). Such migration behavior

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enlarges estrogen-contaminated areas and threatens the quality of surface water (Jenkins et

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al., 2006; Khanal et al., 2006) and groundwater (Wicks et al., 2004; Swartz et al., 2006). In

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a study conducted in California, steroid estrogen was observed in 86% of pasture surface

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water samples (Kolodziej et al., 2007). Even when influent is processed by sewage

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treatment plants (STPs), concentrations of E2 in effluent can still reach 64 ng L−1 in some

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countries (Ying et al., 2002). 3

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Microbial degradation is a significant channel for reducing the risk of E2

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contamination in aquatic systems. However, the co-existence of other pollution factors

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may influence such biological processes, and antibiotics are a candidate for such a role. As

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broad-spectrum antibiotics, tetracyclines (TCs) are commonly ingested by animals and

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poultry as veterinary drugs. They account for 14% of the total amount of antibiotics used

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for treating human and animal infections (Xu et al., 2007; Zhao et al., 2010). Owing to

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surface runoff, percolation, and riparian filtration, ng L−1–µg L−1 levels of TCs were

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frequently detected in treated wastewater effluent, surface water, and groundwater

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(Karthikeyan and Meyer, 2006; Holm et al., 1995). Oxytetracycline (OTC) alone can

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reach 68 µg L−1 in stream waters near livestock farms (Matsui et al., 2008). Moreover,

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OTC concentration with a record high of 0.36 mg L−1 was reported in the Wangyang River,

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a typical river receiving sewage discharges in north China (Jiang et al., 2014).

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The combined pollution caused by E2 and TCs in aquatic ecosystems may be

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universally present in surface aqueous solutions, not only increasing the risk of

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contamination by antibiotic resistance genes but also intensifying the disruption of the

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endocrine systems of aquatic organisms. Because TCs modulate the expression of genes

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related to steroidogenesis to promote the production of hormones (Gracia et al., 2007),

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they show similar endocrine-disrupting effects to E2. Based on limited literature, the

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presence of antibiotics was confirmed to increase the persistence of natural estrogens,

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although the exact mechanism is unclear (He et al., 2019). There is no lack of studies on

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the biodegradation of E2 (Kurisu et al., 2010; Yu et al., 2007; Chen et al., 2017;

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Hashimoto et al., 2010; Fujii et al., 2002) and TCs (Leng et al., 2016; Qi et al., 2019), but 4

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how TCs in aqueous solutions disturb the degradation of E2 in terms of their effect on the

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metabolic pathway is unknown.

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In view of these problems, this study used Novosphingobium sp. ES2-1, an efficient

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estrogen-degrading bacterium isolated from activated sludge in a domestic STP by our

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laboratory, to comprehensively explore the degradation efficiency and pathway of E2 by

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strain ES2-1 in the presence of TCs. By analyzing the changes of E2 metabolites, the

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impacts of TCs on E2 metabolic mechanism were determined, thereby revealing the

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problems that might arise when functional microorganisms facing compound

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contamination. Besides, the preliminary application of strain ES2-1 in natural water

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further reflected the ability of the degrader in removing target estrogens from oligotrophic

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environment contaminated with general-level TCs. This study provided a novel

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perspective for comprehensively evaluating the risk of estrogen pollution.

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2. Materials and methods

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2.1. Chemicals, media, and strain

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E2 (≥ 98%), estrone (E1; ≥ 98%), and estriol (E3; ≥ 98%) were purchased from

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Sigma-Aldrich (St. Louis, MO, USA). TCs including tetracycline (TC) and OTC (≥ 98%)

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were purchased from J & K Co. (Beijing, China). Table S1 summarized the physical

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properties of the estrogens and TCs. The high-performance liquid chromatography

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(HPLC)-grade methanol and acetonitrile used for estrogen detection were purchased from

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TEDIA (Fairfield, OH, USA). Each type of estrogen was stocked in analytical pure

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acetone (HUSHI, Shanghai, China) to create working solutions with final concentrations 5

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of 5 mg mL−1.

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Mineral salt medium (MSM) (pH 7.0 ± 0.2) and R2A liquid medium used for

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biodegradation experiments and strain propagation were prepared as our previously study

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described (Li et al., 2017); Detailed compositions of the two media were presented in the

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Supplementary Information. MSM supplemented with E2 was named E2 mineral salt

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medium, which was prepared by adding MSM after the acetone used to dissolve estrogen

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was volatilized by sterile air in a sterilized Erlenmeyer flask.

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Novosphingobium sp. ES2-1 has a short-rod shape with no flagellum, and a single

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colony is yellow. It is capable of degrading E1, E2, E3, 4-hydroxyestrone (4-OH-E1), and

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testosterone.

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2.2. Construction of the E2 biodegradation system

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The biodegradation system was constructed in a 20-mL aqueous system comprising

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20 mL sterilized MSM containing 20 mg L−1 E2 and a 5% (v/v) cell suspension of strain

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ES2-1. The cell suspension was prepared as follows: A single colony of strain ES2-1 was

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propagated in R2A liquid medium for 18 h, and then the propagation solution was

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centrifuged at 8000 rpm for 5 min. Next, the supernatant was discarded, and the pellet was

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washed three times with fresh sterilized MSM and then re-suspended in fresh MSM to a

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final optical density at 600 nm (OD600 nm) of 1.0.

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2.3. E2 biodegradation under different TCs concentrations

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To investigate the biodegradation efficiency of E2 by strain ES2-1 under different

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TCs concentrations, 0, 0.1, 0.5, 1, 5, or 10 mg L−1, TC or OTC was added into the

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degradation system, and the mixed system was cultured in a thermostatic oscillation 6

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incubator (30 °C, 150 rpm) for 7 d. After the reaction, residual concentrations of E2 and

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TCs in the degradation system were detected by HPLC equipped with ultraviolet light

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detection (HPLC-UV, Shimadzu LC-20AT, Tokyo, Japan) (conditions detailed below). The

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cell counts of strain ES2-1 were measured using the plate counting method. Treatments

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without strain ES2-1 were set as the control group. All experiments were performed in

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triplicate.

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2.4. Degradation kinetics of E2 and TCs

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0.5 and 8 mg L−1 were set as the representative concentrations of low- or high-level

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TCs, to investigate their effects on E2 biodegradation kinetics, respectively. The

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biodegradation of E2 in the absence of TCs was set as the negative control. For the

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treatments, 0.5, or 8 mg L−1 of TC or OTC were added to the E2 degradation system, and

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the mixture was cultured at 30 °C and 150 rpm for 7 d. Residual E2 and cell counts were

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measured every 24 h by HPLC-UV and plate counting methods, respectively,

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accompanied by the detection of residual TC and OTC (methods detailed below). The

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degradation kinetics of E2 and TCs were both fitted to the pseudo-first-order dynamics

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model. Non-inoculated treatments were set as the blank control. All experiments were

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performed in triplicate.

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2.5. Detection of E2 metabolites

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Sample pretreatments were conducted as follows: 20-mL biodegradation solutions

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were exposed to 40 kHz ultrasonic treatment for 30 min and then desalinated by passing

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through a methanol-preactivated Cleanert polar-enhanced polymer column (500 mg, 6 mL,

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Bonna-Agela Technologies, Tianjing, China) at a flow velocity of 2 mL min−1. The 7

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column was eluted with 5 mL ultra-pure water before eluting the metabolites with 10 mL

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methanol, then the eluent was dried with high purity nitrogen (HPN; 99.999%). The dried

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eluent was re-dissolved with 1 mL HPLC-grade methanol and filtered through a 0.22-µm

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polytetrafluoroethylene filter for later analysis.

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The samples were analyzed by ultra-high-performance liquid chromatography

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coupled with high-resolution mass spectrometry (UPLC-HRMS) (G2-XS QTof, Waters,

159

[Milford, USA]). First, 2 µL of solution was injected into the UPLC column (2.1 × 100

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mm ACQUITY UPLC BEH C18 column containing 1.7-µm particles) at a flow rate of 0.4

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mL min−1. Buffer A was composed of 0.1% formic acid in water; buffer B was composed

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of 0.1% formic acid in methanol. The gradient was 5% buffer B for 0.5 min, 5–95% buffer

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B for 11 min, and 95% buffer B for 2 min. HRMS was performed using an electrospray

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source in positive ion mode with MS acquisition mode, with the selected mass ranging

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from 50 to 1200 m/z. The lock mass option was enabled using leucine-enkephalin (m/z

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556.2771) for recalibration. The ionization parameters were set as follows: sample cone =

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40 V; capillary voltage = 3.0 kV; desolvation gas temperature = 400 °C; source

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temperature = 120 °C. Data were acquired and processed using Masslynx 4.1.

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2.6. Removal of estrogens from natural water by strain ES2-1

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Water samples were collected from three natural aquatic systems flowing through

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Nanjing, China, including Qinhuai River, Yangtze River, and Xuanwu Lake. Cell

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suspensions (10%, v/v) of strain ES2-1 were inoculated into the three water samples (200

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mL), respectively, and the mixtures were cultivated in a rotary shaker at 30 °C and 150

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rpm for 7 d. Initial and residual concentrations of estrogens before and after inoculation 8

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with strain ES2-1 were detected by HPLC coupled with a fluorescence detector

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(HPLV-FLD) (methods detailed below), and the removal rate was calculated as follows:

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Removal rate (%) =

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2.7. Analysis of estrogens in culture solution and natural waters

    

× 100

(1)

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Detection of estrogens in culture solution was conducted as previously described (Li

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et al., 2017). Briefly, an equal volume of methanol (20 mL) was added into the

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degradation system to extract the estrogens, and the mixture was subjected to

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ultrasound-enhanced dissolution (40 kHz, 30 min), then filtered through a 0.22-µm

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polytetrafluoroethylene filter, and estrogen concentrations were detected by HPLC-UV

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under the following conditions: a 20-µL sample was injected into the pump to flow

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through an Inertsil ODS-SP-C18 column (250 × 4.6 mm; 5 µm) at a flow rate of 1 mL

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min−1; the mobile phase consisted of acetonitrile and ultra-pure water at a ratio of 70/30

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(v/v); the column temperature was maintained at 40 °C, and the UV detection wavelength

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was set to 280 nm. The limit of detection (LOD, S/N = 3) and limit of quantitation (LOQ,

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S/N = 10) of E3, E2, and E1 were 0.079 and 0.263 mg L−1, 0.062 and 0.207 mg L−1, as

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well as 0.098 and 0.327 mg L−1, respectively.

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Concentrations of estrogens in natural water samples were detected using solid-phase

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extraction (SPE) coupled to HPLC-FLD (Liu et al., 2018). A 200-mL water sample was

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pumped through a C18 SPE column at a flow rate of 5 mL min−1 to purify and concentrate

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the estrogen, and then the column was eluted with 10 mL methanol after washing with 5

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mL water. The eluent was then dried with HPN and brought to a volume of 1 mL with

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methanol.

Impurities

were

removed

by 9

filtration

through

a

0.22-µm

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polytetrafluoroethylene filter. The mobile phase comprised acetonitrile, methanol, and

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water at a ratio of 30/25/45 (v/v/v); the column temperature was set to 40 °C, and 20-µL

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samples were subjected to HPLC-FLD with excitation and emission wavelengths of 280

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nm and 310 nm, respectively. The LOD (S/N = 3) and LOQ (S/N = 10) of E3, E2, and E1

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were 0.18 and 0.60 µg L−1, 0.19 and 0.63 µg L−1, as well as 0.27 and 0.90 µg L−1,

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respectively.

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2.8. Analysis of TCs in culture solution and natural waters

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Sample preparation steps prior to analysis of TCs in culture solution were the same as

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those for estrogen analysis, expect for differences in the HPLC-UV conditions. An Inertsil

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ODS-SP-C18 column (250 × 4.6 mm; 5 µm) was used; the mobile phase, which was

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composed of acetonitrile/methanol/ultrapure water containing 0.1% formic acid (v/v) at a

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ratio of 20/20/60 (v/v/v), flowed into the column at a rate of 0.4 mL min−1. The column

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temperature and UV detection wavelength were maintained at 35 °C and 355 nm,

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respectively; the injection volume was 20 µL (Liu et al., 2016). The LOD (S/N = 3) and

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LOQ (S/N = 10) of TC and OTC were 0.071 and 0.237 mg L−1, as well as 0.064 and 0.213

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mg L−1, respectively.

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Extraction of TCs from natural water samples was identical to estrogen extraction

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from water samples. A 200-mL water sample was pumped through a C18 SPE column at a

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rate of 5 mL min−1, then the column was eluted with 10 mL methanol after being washed

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with 5 mL water. The eluent was dried with HPN and re-dissolved in 0.5 mL HPLC-grade

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methanol. HPLC-UV conditions for TC detection in water were the same as those in

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culture solution. 10

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2.9. Statistical analysis

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All experiments were conducted in triplicate. All data were processed with Microsoft

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Excel 2016 and Origin 8.5. Each data point in figures and tables represents an average

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value. Standard deviations (SD) in parallel samples are shown in figures as error bars.

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Data comparisons between groups were analyzed by one-way analysis of variance

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(ANOVA) followed by Duncan’s tests at a significance level of p ≤ 0.01. Statistical

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analyses were performed using SPSS 20.0.

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3. Results

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3.1. E2 Biodegradation efficiency of strain ES2-1 in aqueous solution with various

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TCs concentrations

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As shown in Figure 1a, 96% of E2 (20 mg L−1) was degraded by strain ES2-1 in the

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presence of 0.1 mg L−1 TC after 7 d, similar to the degradation efficiency of the 0 mg L−1

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TC control, indicating that TC concentrations ≤ 0.1 mg L−1 had a negligible effect on E2

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biodegradation. As TC concentration increased, biodegradation was suppressed in a

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concentration-dependent manner. When the TC concentration reached 10 mg L−1, there

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was no significant difference in residual E2 between the experimental group (with ES2-1)

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and the control (without ES2-1). Meanwhile, the bacterial biomass was also reduced with

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increasing TC concentrations (Figure 1b). Notably, ~60% of TC remained in the reaction

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systems regardless of the initial concentrations. Also, the residual TC in the control group

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was similar to that in experimental groups.

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Similar to TC, E2 biodegradation was also completely suppressed by OTC at 10 mg 11

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L−1; OTC levels below 0.1 mg L−1 exhibited a negligible effect on E2 biodegradation

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(Figure 1c). With regard to the depletion of antibiotics, only about half of the OTC

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remained after 7 d, revealing that OTC might be more susceptible to loss than TC when

244

compared with TC (Figure 1d). Nevertheless, cell growth arrest by OTC was still greater

245

than that by TC. For instance, as OTC at concentration level of 10 mg L−1, the bacterial

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biomass was barely a hundredth that of equal-level TC (Figure 1d and 1b).

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3.2. Biodegradation kinetics of E2 by strain ES2-1 in aqueous solution with TCs

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Based on the results of previous experiments, 0.5 and 8 mg L−1 TCs were selected as

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the representative concentrations of low- and high-level TCs to investigate their impacts

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on E2 biodegradation kinetics, respectively. In the absence of TCs, more than 90% of E2

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was degraded with time in an exponential decay way due to the microbial degradation,

252

accompanied by a muted increase in the biomass of strain ES2-1 (Figure 2a). When the

253

reaction system was supplemented with 0.5 mg L−1 TC, about 85% of E2 was degraded by

254

strain ES2-1 in the same decay manner, whereas the biomass of strain ES2-1 slightly

255

decreased prior to 4 d and recovered thereafter (Figure 2b). The decay manner by which

256

strain ES2-1 degrade E2 under 0.5 mg L−1 OTC stress resembled those of under 0.5 mg

257

L−1 TC, and the biomass of strain ES2-1 also decreased slightly and increased

258

subsequently (Figure 2c).

259

Both the E2 biodegradation rate and the propagation of strain ES2-1 were distinctly

260

hampered at elevated TCs concentrations. Only 50% of E2 was gradually degraded in a

261

weak exponential manner when supplied with 8 mg L−1 TC, and the biodegradation rate

262

was extremely low at the later reaction stage (Figure 2d). The effect of 8 mg L−1 OTC on 12

263

the E2 biodegradation rate was analogous to that of TC at equal level but slightly more

264

severe in cell growth repression. The growth of strain ES2-1 was severely restrained from

265

1 d, and no cell growth was observed thereafter, the bacterial biomass was only a quarter

266

of those with equal concentration of TC (Figure 2e).

267

It was clear from the equation fitted with the pseudo-first-order kinetics model that

268

E2 biodegradation rate constant (kbd) decreased from 0.33 (without TCs) to 0.25 and 0.23

269

d−1 in the presence of 0.5 mg L−1 TC and OTC, respectively; and kbd even fell to 0.09 and

270

0.08 d−1 when TC and OTC increased to 8 mg L−1, respectively (Figure 2f). Notably, the

271

regression coefficient (R2) deviated more pronouncedly from the desired value (1.0000) in

272

the presence of 8 mg L−1 as compared to 0.5 mg L−1 TCs. As indicated, R2 reached 0.9742

273

and 0.9467 under TC and OTC concentrations of 0.5 mg L−1, whereas it significantly

274

reduced to 0.8166 and 0.8284 when TC and OTC set as 8 mg L−1, respectively. All the

275

results meant that biodegradation period was extended with the addition of TCs, resulting

276

in the deviation between biodegradation kinetics curves and the pseudo-first-order kinetics

277

model, especially under the action of high-level TCs. Second, OTC exerted a more potent

278

effect on inhibiting E2 biodegradation than did equal-level TC.

279

3.3. Degradation of TCs

280

The decrease in TCs was observed during the biodegradation of E2. 0.5 mg L−1 TC

281

was reduced regularly with time, and the degradation curves in the experimental and

282

control groups nearly overlapped (Figure 3a); Furthermore, the degradation behavior of 8

283

mg L−1 TC bore a remarkable resemblance to those of 0.5 mg L−1 TC (Figure 3b). This

284

suggested that the decrease in TC was irrelevant with microbial degradation and its initial 13

285

concentration. Likewise, the decay of OTC was also unrelated to those two factors (Figure

286

3c and 3d). The only difference was that OTC exhibited a more visibly exponential decay

287

than did equal-level TC.

288

Pseudo-first-order kinetic curve was also used to describe the degradation behavior of

289

TCs. It was clearly shown in Figure 3e that pseudo-first-order kinetic was propitious to

290

describe such behavior, with the R2 ranging from 0.953 to 0.991. Regardless of the

291

presence or the absence of microorganism, or the different initial concentrations of TC, the

292

degradation rate constant (kd) of TC only fluctuated within the error range and kept at

293

0.078–0.092 d−1. By analogy, the kd of OTC was obtained in the range of 0.131–0.144 d−1.

294

However, the kd of OTC was obviously higher than equal concentration of TC. For

295

example, the kd of TC at 0.5 mg L−1 in the presence and absence of strain ES2-1 were

296

0.091 and 0.092 d−1, while that of OTC at 0.5 mg L−1 reached 0.158 and 0.145 d−1,

297

respectively. As the results the reflected by kinetic parameters, the degradation rate of TCs

298

had no correction with their initial concentrations, and that OTC was degraded faster than

299

equal-level TC.

300

3.4. Metabolites and metabolic pathways of E2 by strain ES2-1 in aqueous solution

301

with TCs

302

In aqueous solution with TCs, eleven E2 metabolites belonging to three possible

303

degradation pathways were identified by UPLC-HRMS (Table 1); the corresponding

304

spectra of these metabolites were shown in Figure S1. m/z [M+H]+ found at 271.1690

305

(product P1, E1), m/z 287.1633 (product P2, 4-OH-E1), m/z 319.1535 (product P4), and

306

m/z 369.1560 (product P7) proved the existence of pathway 1, through which E2 was 14

307

cleaved until the complete opening of the steroidal nucleus. Besides, owing to the m/z

308

[M+H]+ found at 272.1635 (product N1), 304.1575 (products N3), and 308.1545 (products

309

N5), it was proved that there existed a pathway 2 differing from pathway 1, through which

310

part of phenolic ring-opening products could condense with NH3 to form the pyridine

311

derivatives, accompanied by decarbonylation at C4. On the contrary, decarbonylation

312

might not occur as the identification of products Ns1 (pyridinestrone acid) through Ns4,

313

with m/z values found at 300.1573, 316.1525, 332.1528, and 350.1649, respectively.

314

Consequently, pathway 3 was proposed, through which all metabolites performed a special

315

structure of pyridine carboxylic acid.

316

E2 ring cleavage was influenced by TCs, especially at high-level concentrations, as

317

reflected by the decreased abundance of ring-opening metabolites (Figure 4). More than

318

ten products were observed in the reaction system exposed to TC at 0.5 mg L−1 (Figure 4a),

319

whereas at 8 mg L−1, the abundance of metabolites was lower, but an obvious amassing of

320

products P2, N1, and Ns1 occurred (Figure 4b). Results illustrated that increased

321

concentrations of TC dramatically restrained the production of oxidative ring-opening

322

products but urged various pyridine derivatives to accumulate. The effect of OTC on E2

323

metabolism resembled that of TC but was slightly more severe. Specifically, total peak

324

area of the eleven products in the groups treated with 0.5 mg L−1 OTC was lower than that

325

of groups with 0.5 mg L−1 TC, and it was barely weakened over time (Figure 4c); When

326

OTC raised to 8 mg L−1, ring-opening products such as P4, N3, and Ns2–Ns4 were

327

scarcely detected, whereas P2 and Ns1 accumulated significantly, even when compared to

328

their accumulation under an equal concentration of TC (Figure 4d). These findings 15

329

suggested that, first, the suppression of E2 biodegradation by OTC was more difficult to

330

diminish with time than that of TC. Second, enhancing the concentration of TCs could

331

remarkably restrain E2 ring opening, but promoting the condensation of some phenol ring

332

opening products with NH3 in the reaction system, leading to the accumulation of pyridine

333

derivatives.

334

3.5. Removal of estrogens by strain ES2-1 in natural water containing TCs

335

To assess the removal ability of E2 by strain ES2-1 in realistic aquatic systems,

336

laboratory experiments based on oligotrophic environments containing TCs were designed.

337

Water samples were collected from three water systems flowing through Nanjing,

338

including the Qinhuai River, Yangtze River, and Xuanwu Lake. The basic information of

339

water samples was listed in Table S2. TC and OTC concentrations in the three water

340

samples ranged from 0.20 to 3.41 µg L−1, and from 0.72 to 3.85 µg L−1, respectively (Table

341

2a). Concentrations of E3, E2, and E1 in the three water samples were 0.77–7.77 µg L−1,

342

0.17–0.88 µg L−1, and non-detected (ND)–7.80 µg L−1, respectively. 7 d later, residual E3,

343

E2, and E1 concentrations in the three water samples without inoculation with strain

344

ES2-1 were slightly decreased, by 10.89–20.35%, 10.45–16.80%, and 13.42–19.25%,

345

respectively. When the water samples were treated with strain ES2-1 for 7 d, residual E3,

346

E2, and E1 concentrations were reduced to ND–0.61 µg L−1, ND–0.08 µg L−1, and 0–0.15

347

µg L−1, with removal rates of 72.38–100%, 52.10–100%, and 98.06–100%, respectively

348

(Table 2b). The results showed that the strain could still degrade estrogens at

349

environment-related concentrations in natural oligotrophic waters that might contain TCs.

350 16

351

4. Discussion

352

For a water system potentially polluted by organic pollutants, microbial degradation

353

is critical in reducing the risk of such pollution. However, studies tend to focus on the

354

biodegradation of one type of pollutant while ignoring the possible impacts brought about

355

by coexisting pollutants. Estrogens and antibiotics are two groups of pollutants of concern

356

in large-scale farming. Therefore, this study investigated the effects of the presence of TCs

357

(representing antibiotics) on the biodegradation mechanisms of E2 (representing

358

estrogens).

359

2 mg L−1 was the minimum inhibitory concentration (MIC) of TCs on

360

Novosphingobium sp. ES2-1, the degrader of interest in this study (Figure S2). E2

361

biodegradation under the actions of low-level (≤ 2 mg L−1, represented by 0.5 mg L−1) TCs

362

was quite different from that of under high concentrations (≥ 2 mg L−1, represented by 8

363

mg L−1). Cell growth was slightly disturbed by low-level TCs but recovered in later

364

reaction stage, and the degradation behavior could still be described using

365

pseudo-first-order kinetics model. This was because, first, further reduction in low-level

366

TCs relieved the already low toxicity to functional strain. Second, low-level antibiotics

367

belonged to concentrations that lower than the MIC (sub-MIC) and could select resistance

368

strains. By duplications of microorganisms, a pre-existing weak resistance phenotype

369

could be amplified, thereby increasing the probability of bacterial survival and growth in

370

the presence of antibiotics (Andersson and Hughes, 2014). By contrast, bacteria exposed

371

to high-level antibiotic were harder to survive. Hydrolysis alone was insufficient to rapidly

372

reduce the toxicity of high-level TCs on microorganisms to safe levels. A slowly- or 17

373

non-growing state was therefore needed for bacteria to be alive with such level of

374

antibiotics (Allison et al., 2011). Early in the reaction stage, high-level TCs severely

375

suppressed the cell growth or even killed the population. A great restriction on the initial

376

biotransformation rate was thus formed, thus making pseudo-first-order kinetic unsuitable

377

for describing the biodegradation behavior under high-level TCs.

378

The reduction in TCs occurred simultaneously with E2 biodegradation, and its

379

manner of decay could also be described by pseudo-first-order kinetic model. However,

380

the decrease of TCs had no connection with biodegradation or photodegradation, making

381

hydrolysis probably the best explanation for such decline. Xuan et al. (2009) also reported

382

the hydrolysis of TCs in deionized water and fitted the hydrolysis curve of OTC with the

383

same model, and hydrolysis rate constant (kh) was obtained in the range of 0.094–0.106

384

d−1 at 25 °C; While the kh were within the range of 0.134–0.142 d−1 in our results but at

385

30 °C, the difference could be attribute to the reaction temperature. As TCs are heat-labile,

386

high temperatures accelerate the hydrolysis rate (Leffler and Grunwald, 1963). As

387

amphoteric molecules, the kh of TCs is also related to pH value of the reaction system.

388

TCs are easily hydrolyzed under neutral, slightly acidic, and alkaline aqueous conditions,

389

with aqueous half-lives ranging from 20 h to 50 d (Doi et al., 2000; Pouliquen et al., 2007;

390

Loftin et al., 2008; Xuan et al., 2009; Kang et al., 2010). The pH value established in this

391

study was close to neutral and fluctuated little during the reaction, which satisfied the

392

optimal requirement of TCs hydrolyzation.

393

The presence of TCs reduced not only the degradation efficiency but also the

394

abundance of ring-opening metabolites. The proposed pathways were shown in Figure 5. 18

395

Product P3 that probably generated via the 4,5-seco pathway (Chen et al., 2017) was

396

suggested to be the metabolite between P2 and P4, but the m/z evidence supporting its

397

generation in the presence of TCs was lacking. Product P5 and P6 were both downstream

398

of P4, where P5 might be produced through the 9,10-seco pathway (Horinouchi et al.,

399

2004; Sih et al., 1965), whereas their existences were also failed to be supported by mass

400

evidence when TCs were added. Similarly, the production of the proposed derivate

401

products N2 and N4 in the presence of TCs was not well supported. A particularly

402

interesting case was the accumulation of some pyridine derivatives. Ns1 was documented

403

in previous reports, and its formation mechanism was ascribed to the condensation of

404

phenol ring-cleavage products with ammonium ions (Dagley et al., 1960; Coombe et al.,

405

1966). The presence of TCs, especially at raised concentrations, significantly hindered the

406

ring-opening process but improved the contact probability between some phenolic

407

ring-opening products and ammonium ions, enabling E2 metabolism to linger in the stage

408

of condensation reaction. Moreover, the abundances of metabolites generated through

409

pathway 2 were obviously lower than those in pathway 3. To our knowledge, these two

410

pathways differ only in the loss of a CO2 molecule, which might be caused by certain

411

decarboxylase secreted by strain ES2-1 during the condensation reaction. It is well known

412

that TCs can weaken the interaction between ribosomes and tRNAs, impeding protein

413

synthesis (Epe et al. 1987), and the decarboxylase might be involved. This hypothesis

414

deserves further investigation. Remarkably, both N1 and Ns1 shared the same steroid

415

nucleus, yet nitrogen was introduced in such nucleus, so it remained to be seen whether

416

these derivatives still performed estrogenic activity. Furthermore, N1 and Ns1 might be 19

417

produced in any aqueous solution contained ammonium ions, the presence of antibiotics

418

just facilitated their accumulation. Hence, it may be an emerging factor deserved to be

419

considered in assessing the ecological risk.

420

Concentrations of natural estrogens in real water bodies are usually at ng L−1–µg L−1

421

levels (Furuichi et al., 2004; Laganà et al., 2004; Lei et al., 2009; Kolodziej et al., 2004;

422

Sarmah et al., 2006; Shappell et al., 2007; Velicu and Suri, 2009), the levels of estrogens

423

detected in this study generally coincided with those reported previously. Although TCs at

424

µg L−1 levels were detected in this study, it remained within the reasonable concentration

425

range. Nevertheless, such level of antibiotics might enable microbes to generate and

426

maintain resistance. More consideration could be given to the two-sidedness of functional

427

degrading bacteria developing drug resistance when assessing the ecological risks of

428

combined contamination. In oligotrophic water bodies containing TCs, the effective

429

degradation of estrogens by functional strain was primarily to blame the instability of TCs

430

due to hydrolysis, etc., causing the impermanence of their toxic effects on microorganisms.

431

Second, the oligotrophic environments starved functional microorganisms, obliging them

432

to utilize more organic pollutants to meet their survival needs (Marshall et al., 2000).

433

Third, strain ES2-1 was better adapted to low-nutrition conditions than those

434

microorganisms survived in nutritious circumstances, as it was isolated from a STP

435

containing various pollution factors.

436 437 438

5. Conclusions Levels of TCs have an important influence on the degradation of E2 by functional 20

439

microorganisms. The effects of TCs at sub-MIC on E2 biodegradation were quite different

440

from those at concentrations above MIC. The presence of TCs reduced the degradation

441

rate and integrity of E2 by functional bacteria through interfering with the subsequent

442

ring-opening process, accompanied by the condensation of various phenolic ring-opening

443

products with ammonium ions and finally causing E2 metabolism to linger in the

444

condensation stage. Notwithstanding the levels of TCs in the natural aqueous environment

445

were generally low, which was insufficient to cause significant biodegradation interference,

446

water systems that receive high-level antibiotics might still face such uncertain risks. More

447

studies can focus on assessing the estrogenic activity of those derivatives, as well as

448

shielding functional microorganisms from the injury of TCs during biodegradation.

449 450

Author contributions

451

S.L., J.L., and W.L. conceived the project and designed the experiments. S.L., J.L,

452

and K.S. performed the experiments. S.L. and Z.Y. performed statistical analyses. S.L.,

453

K.S., Z.Y., and W. L. wrote the paper. All authors edited and approved the final

454

manuscript.

455 456

Declaration of Interest Statement

457

The authors declare that they have no known competing financial interests or

458

personal relationships that could have appeared to influence the work reported in this

459

paper.

460 21

461

Acknowledgments

462

This work was financially supported by the National Natural Science Foundation of

463

China (41977121, 41771523), and the National Key Research and Development Program

464

of China (2016YFD0800203).

465 466

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29

616

Legends of figures and tables

617

Figure 1. E2 Biodegradation efficiency of strain ES2-1 in aqueous solution with TCs at

618

various concentrations. (a) Influence of different TC concentrations on E2

619

biodegradation efficiency; (b) Residual TC and cell counts of strain ES2-1 in the

620

degradation system after 7 d; (c) Influence of different OTC concentrations on E2

621

biodegradation efficiency; (d) Residual OTC and biomass of strain ES2-1. * p < 0.05,

622

** p < 0.01, Duncan’s test, control (without ES2-1) versus experimental groups (with

623

ES2-1) at different concentrations of TCs.

624

Figure 2. E2 biodegradation kinetics in the presence of low- or high-level TCs represented

625

by 0.5 and 8 mg L−1 TCs, respectively. (a) E2 biodegradation kinetics without TCs;

626

E2 biodegradation kinetics under the action of low-level TC (b) or OTC (c); E2

627

biodegradation kinetics under the action of high-level TC (d) or OTC (e); (f) E2

628

biodegradation kinetic parameters obtained by fitting with the pseudo-first-order

629

kinetic model. * p < 0.05, ** p < 0.01, Duncan’s test, groups with 0.5 mg L−1 TCs

630

versus groups with 8 mg L−1 TCs.

631

Figure 3. Degradation kinetics of TCs during the biodegradation of E2. Degradation

632

kinetics curves of 0.5 (a) and 8 (b) mg L−1 TC; Degradation kinetics curves of 0.5 (c)

633

and 8 (d) mg L−1 OTC; (e) Degradation kinetics parameters of TCs with different

634

initial concentrations after fitting with the pseudo-first-order dynamic model. * p <

635

0.05, ** p < 0.01, Duncan’s test, groups treated by TC versus groups treated by OTC.

636

Figure 4. Abundances of E2 metabolites in the presence of TCs. Products P1, P2, P4, and

637

P7 were metabolites generated through pathway 1; products N1, N3, and N5 were 30

638 639 640

produced via pathway 2; products Ns1 through Ns4 were produced via pathway 3. Figure 5. Proposed biodegradation pathway of E2 by strain ES2-1 in the presence of TCs. The name of the metabolite was given next to the structural formula.

641

Table 1. UPLC-HRMS data of E2 metabolites.

642

Table 2. Removal of estrogens from natural water samples containing TCs by strain ES2-1.

643

(a) Initial and residual (after 7 d) concentrations of TCs in three water samples; (b)

644

Initial and residual (after 7 d) concentrations of E3, E2, E1 in different inoculation

645

treatment groups.

31

646 647

Figure 1

32

648 649

Figure 2

33

650 651

Figure 3

34

652 653

Figure 4

35

654 655

Figure 5

36

656

Table 1 Chemical

Retention time

m/z

formula

(min)

([M+H]+)

C18H24O2

5.76

273.1833

-0.0022

Substrate

C18H22O2

6.48

271.1690

-0.0018

Yu et al., 2007

C18H22O3

5.76

287.1633

-0.0014

Chen et al., 2017

P4

C18H22O5

3.03

319.1535

-0.0010

Coombe et al., 1966

P7

C18H24O8

1.88

369.1560

0.0011

This study

N1

C17H21NO2

1.68

272.1635

-0.0016

This study

N3

C17H21NO4

3.42

304.1575

0.0026

This study

N5

C16H21NO5

2.69

308.1545

0.0047

This study

Ns1

C18H21NO3

2.68

300.1573

-0.0027

Name

E2

Mass error

Predicted chemical structure

References

P1

Pathway 2

Pathway 1

(E1) P2 (4-OH-E1)

Chen et al., 2017 Coombe et al., 1966

Pathway 3

O

Ns2

C18H21NO4

2.12

316.1525

-0.0024

Sih et al., 1965 O HOOC

N

O

Ns3

C18H21NO5

3.68

332.1528

0.0030

This study O HOOC

Ns4

C18H23NO6

2.51

350.1649

657

37

0.0045

O

N

This study

658

Table 2

659

(a) TCs in natural water Initial concentrations (µg L−1)

Residual concentrations (µg L−1)

Antibiotics

660

1

2

3

1

2

3

TC

0.30 ± 0.1

0.20 ± 0.01

3.41 ± 0.35

0.20 ± 0.07

ND

1.37 ± 0.40

OTC

0.72 ± 0.08

2.09 ± 1.07

3.85 ± 2.01

0.33 ± 0.06

1.16 ± 0.44

1.98 ± 0.78

(b) Removal of estrogens in natural water Initial concentration (µg L−1)

Residual concentration (µg L−1)

Removal rate (%)

Samples E3

E2

E1

E3

1

2.20 ± 0.33

0.17 ± 0.02

ND

1.75 ± 0.13

2

7.77 ± 0.25

0.88 ± 0.05

7.80 ± 0.87

6.42 ± 0.26

3

0.77 ± 0.15

0.45 ± 0.08

0.31 ± 0.06

0.68 ± 0.09

1

2.20 ± 0.33

0.17 ± 0.02

ND

0.61 ± 0.05

2

7.77 ± 0.25

0.88 ± 0.05

7.80 ± 0.87

ND

3

0.77 ± 0.15

0.45 ± 0.08

0.31 ± 0.06

ND

E2

E1

E3

E2

E1

0.1 5 ± 0.03

ND

20.35

10.45



0.78 ± 0.05

6.75 ± 1.17

17.45

11.41

13.42

0.38 ± 0.03

0.25 ± 0.02

10.89

16.80

19.25

0.08 ± 0.04

ND

72.38

52.10



0.01 ± 0.00

0.15 ± 0.03

100

98.65

98.06

ND

ND

100

100

100

Non-inoculated

Inoculated

661

38

Highlights ·E2 biodegradation was significantly disturbed at elevated TCs concentrations. ·TCs hydrolysis happened simultaneously with E2 biodegradation. ·TCs disturbed the cleavage of E2, reducing the abundance of ring-opening products. ·TCs enabled E2 metabolism to linger in the stage of condensing with NH3. ·Strain ES2-1 could effectively remove E2 in natural waters containing TCs.

Author contributions S.L., J.L., and W.L. conceived the project and designed the experiments. S.L., J.L., and K.S. performed the experiments. S.L. and Z.Y. performed statistical analyses. S.L., K.S., Z.Y., and W. L. wrote the paper. All authors edited and approved the final manuscript.

Declaration of interests The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.