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:
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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.
Degradation of 17β-estradiol by Novosphingobium sp. ES2-1 in
aqueous solution contaminated with tetracyclines
Shunyao Lia,1, Juan Liua,1, Kai Sun2, Zhiyao Yang1, Wanting Ling*,1
and Environmental Sciences, Nanjing Agricultural University, Nanjing 210095, China
Prevention, School of Resources and Environment, Anhui Agricultural University, Hefei,
Institute of Organic Contaminant Control and Soil Remediation, College of Resources
Anhui Province Key Laboratory of Farmland Ecological Conservation and Pollution
*Corresponding author: Wanting Ling, Dr/Professor
Address: Weigang Road 1, Nanjing 210095, China
E-mail: [email protected]
These authors contributed equally to this paper.
19 20 21
17β-estradiol (E2) often coexists with tetracyclines (TCs) in wastewater lagoons at
intensive breeding farms, threatening the quality of surrounding water bodies. Microbial
degradation is vital in E2 removal, but it is unclear how TCs affect E2 biodegradation.
This primary study investigated the mechanisms of E2 degradation by Novosphingobium
sp. ES2-1 in the presence of TCs and assessed the removal efficiency of E2 by strain
ES2-1 in natural waters containing TCs. E2 biodegradation was unaffected at TCs
concentrations below 0.1 mg L−1 yet significantly inhibited at TCs above 10 mg L−1. As
elevation of TCs, E2 biodegradation rate constant decreased, and the biodegradation
kinetics equation gradually deviated from the pseudo-first-order dynamics model.
Importantly, the presence of TCs, especially at high-level concentrations, significantly
hindered E2 ring-opening process but promoted the condensation of some phenolic
ring-opening products with NH3, thereby increasing the abundance of pyridine derivatives,
which were difficult to decompose over time. Additionally, strain ES2-1 could remove
52.1–100% of nature estrogens in TCs-contaminated natural waters within 7 d. Results
revealed the mechanisms of TCs in E2 biodegradation and the performance of a functional
strain in estrogen removal in realistic TCs-contaminated aqueous solution.
Keywords: Antibiotics; Estrogens; Degradation; Sphingomonad; Combined pollution
Capsule: TCs may inhibit the E2 degradation by strain ES2-1, and disturb the cleavage of
E2, reducing the abundance of ring-opening products.
17β-estradiol (E2) is an estrogen of priority concern due to its high endocrine activity
(Bradley et al., 2009). If allowed to accumulate in the ecosystem and enter the human food
chain, E2 can not only disturb the sex ratio balance of aquatic wildlife but induce human
cancer and increase the probability of viral infections (Wocławek-Potocka et al., 2013;
Arnold et al., 2014; Tetreault et al., 2011; Rose et al., 2013; Yang et al., 2018). Both the
US Environmental Protection Agency and the European Commission have included
estrogens in regulatory lists due to their potential side effects (US-EPA, 2016; EE
Commission, 2013). Livestock waste is a major source of estrogens. It was estimated that
estrogens discharged by livestock in the United States and European Union reached
83,000 kg per year (Adeel et al., 2017). The UK average annual emission of E2 in 2013
reached 570 kg (Ray et al., 2013).
Storage of liquid manure, manure spreading, and surface runoff provide convenient
modes for E2 in livestock waste to enter water systems, leading to high detection
frequency of E2 in surface waters, with levels ranging from ng L−1 to µg L−1 (Beck et al.,
2006; Kjær et al., 2007; Pal et al., 2010; Rocha et al., 2012). Such migration behavior
enlarges estrogen-contaminated areas and threatens the quality of surface water (Jenkins et
al., 2006; Khanal et al., 2006) and groundwater (Wicks et al., 2004; Swartz et al., 2006). In
a study conducted in California, steroid estrogen was observed in 86% of pasture surface
water samples (Kolodziej et al., 2007). Even when influent is processed by sewage
treatment plants (STPs), concentrations of E2 in effluent can still reach 64 ng L−1 in some
countries (Ying et al., 2002). 3
Microbial degradation is a significant channel for reducing the risk of E2
contamination in aquatic systems. However, the co-existence of other pollution factors
may influence such biological processes, and antibiotics are a candidate for such a role. As
broad-spectrum antibiotics, tetracyclines (TCs) are commonly ingested by animals and
poultry as veterinary drugs. They account for 14% of the total amount of antibiotics used
for treating human and animal infections (Xu et al., 2007; Zhao et al., 2010). Owing to
surface runoff, percolation, and riparian filtration, ng L−1–µg L−1 levels of TCs were
frequently detected in treated wastewater effluent, surface water, and groundwater
(Karthikeyan and Meyer, 2006; Holm et al., 1995). Oxytetracycline (OTC) alone can
reach 68 µg L−1 in stream waters near livestock farms (Matsui et al., 2008). Moreover,
OTC concentration with a record high of 0.36 mg L−1 was reported in the Wangyang River,
a typical river receiving sewage discharges in north China (Jiang et al., 2014).
The combined pollution caused by E2 and TCs in aquatic ecosystems may be
universally present in surface aqueous solutions, not only increasing the risk of
contamination by antibiotic resistance genes but also intensifying the disruption of the
endocrine systems of aquatic organisms. Because TCs modulate the expression of genes
related to steroidogenesis to promote the production of hormones (Gracia et al., 2007),
they show similar endocrine-disrupting effects to E2. Based on limited literature, the
presence of antibiotics was confirmed to increase the persistence of natural estrogens,
although the exact mechanism is unclear (He et al., 2019). There is no lack of studies on
the biodegradation of E2 (Kurisu et al., 2010; Yu et al., 2007; Chen et al., 2017;
Hashimoto et al., 2010; Fujii et al., 2002) and TCs (Leng et al., 2016; Qi et al., 2019), but 4
how TCs in aqueous solutions disturb the degradation of E2 in terms of their effect on the
metabolic pathway is unknown.
In view of these problems, this study used Novosphingobium sp. ES2-1, an efficient
estrogen-degrading bacterium isolated from activated sludge in a domestic STP by our
laboratory, to comprehensively explore the degradation efficiency and pathway of E2 by
strain ES2-1 in the presence of TCs. By analyzing the changes of E2 metabolites, the
impacts of TCs on E2 metabolic mechanism were determined, thereby revealing the
problems that might arise when functional microorganisms facing compound
contamination. Besides, the preliminary application of strain ES2-1 in natural water
further reflected the ability of the degrader in removing target estrogens from oligotrophic
environment contaminated with general-level TCs. This study provided a novel
perspective for comprehensively evaluating the risk of estrogen pollution.
2. Materials and methods
2.1. Chemicals, media, and strain
E2 (≥ 98%), estrone (E1; ≥ 98%), and estriol (E3; ≥ 98%) were purchased from
Sigma-Aldrich (St. Louis, MO, USA). TCs including tetracycline (TC) and OTC (≥ 98%)
were purchased from J & K Co. (Beijing, China). Table S1 summarized the physical
properties of the estrogens and TCs. The high-performance liquid chromatography
(HPLC)-grade methanol and acetonitrile used for estrogen detection were purchased from
TEDIA (Fairfield, OH, USA). Each type of estrogen was stocked in analytical pure
acetone (HUSHI, Shanghai, China) to create working solutions with final concentrations 5
of 5 mg mL−1.
Mineral salt medium (MSM) (pH 7.0 ± 0.2) and R2A liquid medium used for
biodegradation experiments and strain propagation were prepared as our previously study
described (Li et al., 2017); Detailed compositions of the two media were presented in the
Supplementary Information. MSM supplemented with E2 was named E2 mineral salt
medium, which was prepared by adding MSM after the acetone used to dissolve estrogen
was volatilized by sterile air in a sterilized Erlenmeyer flask.
Novosphingobium sp. ES2-1 has a short-rod shape with no flagellum, and a single
colony is yellow. It is capable of degrading E1, E2, E3, 4-hydroxyestrone (4-OH-E1), and
2.2. Construction of the E2 biodegradation system
The biodegradation system was constructed in a 20-mL aqueous system comprising
20 mL sterilized MSM containing 20 mg L−1 E2 and a 5% (v/v) cell suspension of strain
ES2-1. The cell suspension was prepared as follows: A single colony of strain ES2-1 was
propagated in R2A liquid medium for 18 h, and then the propagation solution was
centrifuged at 8000 rpm for 5 min. Next, the supernatant was discarded, and the pellet was
washed three times with fresh sterilized MSM and then re-suspended in fresh MSM to a
final optical density at 600 nm (OD600 nm) of 1.0.
2.3. E2 biodegradation under different TCs concentrations
To investigate the biodegradation efficiency of E2 by strain ES2-1 under different
TCs concentrations, 0, 0.1, 0.5, 1, 5, or 10 mg L−1, TC or OTC was added into the
degradation system, and the mixed system was cultured in a thermostatic oscillation 6
incubator (30 °C, 150 rpm) for 7 d. After the reaction, residual concentrations of E2 and
TCs in the degradation system were detected by HPLC equipped with ultraviolet light
detection (HPLC-UV, Shimadzu LC-20AT, Tokyo, Japan) (conditions detailed below). The
cell counts of strain ES2-1 were measured using the plate counting method. Treatments
without strain ES2-1 were set as the control group. All experiments were performed in
2.4. Degradation kinetics of E2 and TCs
0.5 and 8 mg L−1 were set as the representative concentrations of low- or high-level
TCs, to investigate their effects on E2 biodegradation kinetics, respectively. The
biodegradation of E2 in the absence of TCs was set as the negative control. For the
treatments, 0.5, or 8 mg L−1 of TC or OTC were added to the E2 degradation system, and
the mixture was cultured at 30 °C and 150 rpm for 7 d. Residual E2 and cell counts were
measured every 24 h by HPLC-UV and plate counting methods, respectively,
accompanied by the detection of residual TC and OTC (methods detailed below). The
degradation kinetics of E2 and TCs were both fitted to the pseudo-first-order dynamics
model. Non-inoculated treatments were set as the blank control. All experiments were
performed in triplicate.
2.5. Detection of E2 metabolites
Sample pretreatments were conducted as follows: 20-mL biodegradation solutions
were exposed to 40 kHz ultrasonic treatment for 30 min and then desalinated by passing
through a methanol-preactivated Cleanert polar-enhanced polymer column (500 mg, 6 mL,
Bonna-Agela Technologies, Tianjing, China) at a flow velocity of 2 mL min−1. The 7
column was eluted with 5 mL ultra-pure water before eluting the metabolites with 10 mL
methanol, then the eluent was dried with high purity nitrogen (HPN; 99.999%). The dried
eluent was re-dissolved with 1 mL HPLC-grade methanol and filtered through a 0.22-µm
polytetrafluoroethylene filter for later analysis.
The samples were analyzed by ultra-high-performance liquid chromatography
coupled with high-resolution mass spectrometry (UPLC-HRMS) (G2-XS QTof, Waters,
[Milford, USA]). First, 2 µL of solution was injected into the UPLC column (2.1 × 100
mm ACQUITY UPLC BEH C18 column containing 1.7-µm particles) at a flow rate of 0.4
mL min−1. Buffer A was composed of 0.1% formic acid in water; buffer B was composed
of 0.1% formic acid in methanol. The gradient was 5% buffer B for 0.5 min, 5–95% buffer
B for 11 min, and 95% buffer B for 2 min. HRMS was performed using an electrospray
source in positive ion mode with MS acquisition mode, with the selected mass ranging
from 50 to 1200 m/z. The lock mass option was enabled using leucine-enkephalin (m/z
556.2771) for recalibration. The ionization parameters were set as follows: sample cone =
40 V; capillary voltage = 3.0 kV; desolvation gas temperature = 400 °C; source
temperature = 120 °C. Data were acquired and processed using Masslynx 4.1.
2.6. Removal of estrogens from natural water by strain ES2-1
Water samples were collected from three natural aquatic systems flowing through
Nanjing, China, including Qinhuai River, Yangtze River, and Xuanwu Lake. Cell
suspensions (10%, v/v) of strain ES2-1 were inoculated into the three water samples (200
mL), respectively, and the mixtures were cultivated in a rotary shaker at 30 °C and 150
rpm for 7 d. Initial and residual concentrations of estrogens before and after inoculation 8
with strain ES2-1 were detected by HPLC coupled with a fluorescence detector
(HPLV-FLD) (methods detailed below), and the removal rate was calculated as follows:
Removal rate (%) =
2.7. Analysis of estrogens in culture solution and natural waters
Detection of estrogens in culture solution was conducted as previously described (Li
et al., 2017). Briefly, an equal volume of methanol (20 mL) was added into the
degradation system to extract the estrogens, and the mixture was subjected to
ultrasound-enhanced dissolution (40 kHz, 30 min), then filtered through a 0.22-µm
polytetrafluoroethylene filter, and estrogen concentrations were detected by HPLC-UV
under the following conditions: a 20-µL sample was injected into the pump to flow
through an Inertsil ODS-SP-C18 column (250 × 4.6 mm; 5 µm) at a flow rate of 1 mL
min−1; the mobile phase consisted of acetonitrile and ultra-pure water at a ratio of 70/30
(v/v); the column temperature was maintained at 40 °C, and the UV detection wavelength
was set to 280 nm. The limit of detection (LOD, S/N = 3) and limit of quantitation (LOQ,
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
well as 0.098 and 0.327 mg L−1, respectively.
Concentrations of estrogens in natural water samples were detected using solid-phase
extraction (SPE) coupled to HPLC-FLD (Liu et al., 2018). A 200-mL water sample was
pumped through a C18 SPE column at a flow rate of 5 mL min−1 to purify and concentrate
the estrogen, and then the column was eluted with 10 mL methanol after washing with 5
mL water. The eluent was then dried with HPN and brought to a volume of 1 mL with
polytetrafluoroethylene filter. The mobile phase comprised acetonitrile, methanol, and
water at a ratio of 30/25/45 (v/v/v); the column temperature was set to 40 °C, and 20-µL
samples were subjected to HPLC-FLD with excitation and emission wavelengths of 280
nm and 310 nm, respectively. The LOD (S/N = 3) and LOQ (S/N = 10) of E3, E2, and E1
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,
2.8. Analysis of TCs in culture solution and natural waters
Sample preparation steps prior to analysis of TCs in culture solution were the same as
those for estrogen analysis, expect for differences in the HPLC-UV conditions. An Inertsil
ODS-SP-C18 column (250 × 4.6 mm; 5 µm) was used; the mobile phase, which was
composed of acetonitrile/methanol/ultrapure water containing 0.1% formic acid (v/v) at a
ratio of 20/20/60 (v/v/v), flowed into the column at a rate of 0.4 mL min−1. The column
temperature and UV detection wavelength were maintained at 35 °C and 355 nm,
respectively; the injection volume was 20 µL (Liu et al., 2016). The LOD (S/N = 3) and
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
mg L−1, respectively.
Extraction of TCs from natural water samples was identical to estrogen extraction
from water samples. A 200-mL water sample was pumped through a C18 SPE column at a
rate of 5 mL min−1, then the column was eluted with 10 mL methanol after being washed
with 5 mL water. The eluent was dried with HPN and re-dissolved in 0.5 mL HPLC-grade
methanol. HPLC-UV conditions for TC detection in water were the same as those in
culture solution. 10
2.9. Statistical analysis
All experiments were conducted in triplicate. All data were processed with Microsoft
Excel 2016 and Origin 8.5. Each data point in figures and tables represents an average
value. Standard deviations (SD) in parallel samples are shown in figures as error bars.
Data comparisons between groups were analyzed by one-way analysis of variance
(ANOVA) followed by Duncan’s tests at a significance level of p ≤ 0.01. Statistical
analyses were performed using SPSS 20.0.
3.1. E2 Biodegradation efficiency of strain ES2-1 in aqueous solution with various
As shown in Figure 1a, 96% of E2 (20 mg L−1) was degraded by strain ES2-1 in the
presence of 0.1 mg L−1 TC after 7 d, similar to the degradation efficiency of the 0 mg L−1
TC control, indicating that TC concentrations ≤ 0.1 mg L−1 had a negligible effect on E2
biodegradation. As TC concentration increased, biodegradation was suppressed in a
concentration-dependent manner. When the TC concentration reached 10 mg L−1, there
was no significant difference in residual E2 between the experimental group (with ES2-1)
and the control (without ES2-1). Meanwhile, the bacterial biomass was also reduced with
increasing TC concentrations (Figure 1b). Notably, ~60% of TC remained in the reaction
systems regardless of the initial concentrations. Also, the residual TC in the control group
was similar to that in experimental groups.
Similar to TC, E2 biodegradation was also completely suppressed by OTC at 10 mg 11
L−1; OTC levels below 0.1 mg L−1 exhibited a negligible effect on E2 biodegradation
(Figure 1c). With regard to the depletion of antibiotics, only about half of the OTC
remained after 7 d, revealing that OTC might be more susceptible to loss than TC when
compared with TC (Figure 1d). Nevertheless, cell growth arrest by OTC was still greater
than that by TC. For instance, as OTC at concentration level of 10 mg L−1, the bacterial
biomass was barely a hundredth that of equal-level TC (Figure 1d and 1b).
3.2. Biodegradation kinetics of E2 by strain ES2-1 in aqueous solution with TCs
Based on the results of previous experiments, 0.5 and 8 mg L−1 TCs were selected as
the representative concentrations of low- and high-level TCs to investigate their impacts
on E2 biodegradation kinetics, respectively. In the absence of TCs, more than 90% of E2
was degraded with time in an exponential decay way due to the microbial degradation,
accompanied by a muted increase in the biomass of strain ES2-1 (Figure 2a). When the
reaction system was supplemented with 0.5 mg L−1 TC, about 85% of E2 was degraded by
strain ES2-1 in the same decay manner, whereas the biomass of strain ES2-1 slightly
decreased prior to 4 d and recovered thereafter (Figure 2b). The decay manner by which
strain ES2-1 degrade E2 under 0.5 mg L−1 OTC stress resembled those of under 0.5 mg
L−1 TC, and the biomass of strain ES2-1 also decreased slightly and increased
subsequently (Figure 2c).
Both the E2 biodegradation rate and the propagation of strain ES2-1 were distinctly
hampered at elevated TCs concentrations. Only 50% of E2 was gradually degraded in a
weak exponential manner when supplied with 8 mg L−1 TC, and the biodegradation rate
was extremely low at the later reaction stage (Figure 2d). The effect of 8 mg L−1 OTC on 12
the E2 biodegradation rate was analogous to that of TC at equal level but slightly more
severe in cell growth repression. The growth of strain ES2-1 was severely restrained from
1 d, and no cell growth was observed thereafter, the bacterial biomass was only a quarter
of those with equal concentration of TC (Figure 2e).
It was clear from the equation fitted with the pseudo-first-order kinetics model that
E2 biodegradation rate constant (kbd) decreased from 0.33 (without TCs) to 0.25 and 0.23
d−1 in the presence of 0.5 mg L−1 TC and OTC, respectively; and kbd even fell to 0.09 and
0.08 d−1 when TC and OTC increased to 8 mg L−1, respectively (Figure 2f). Notably, the
regression coefficient (R2) deviated more pronouncedly from the desired value (1.0000) in
the presence of 8 mg L−1 as compared to 0.5 mg L−1 TCs. As indicated, R2 reached 0.9742
and 0.9467 under TC and OTC concentrations of 0.5 mg L−1, whereas it significantly
reduced to 0.8166 and 0.8284 when TC and OTC set as 8 mg L−1, respectively. All the
results meant that biodegradation period was extended with the addition of TCs, resulting
in the deviation between biodegradation kinetics curves and the pseudo-first-order kinetics
model, especially under the action of high-level TCs. Second, OTC exerted a more potent
effect on inhibiting E2 biodegradation than did equal-level TC.
3.3. Degradation of TCs
The decrease in TCs was observed during the biodegradation of E2. 0.5 mg L−1 TC
was reduced regularly with time, and the degradation curves in the experimental and
control groups nearly overlapped (Figure 3a); Furthermore, the degradation behavior of 8
mg L−1 TC bore a remarkable resemblance to those of 0.5 mg L−1 TC (Figure 3b). This
suggested that the decrease in TC was irrelevant with microbial degradation and its initial 13
concentration. Likewise, the decay of OTC was also unrelated to those two factors (Figure
3c and 3d). The only difference was that OTC exhibited a more visibly exponential decay
than did equal-level TC.
Pseudo-first-order kinetic curve was also used to describe the degradation behavior of
TCs. It was clearly shown in Figure 3e that pseudo-first-order kinetic was propitious to
describe such behavior, with the R2 ranging from 0.953 to 0.991. Regardless of the
presence or the absence of microorganism, or the different initial concentrations of TC, the
degradation rate constant (kd) of TC only fluctuated within the error range and kept at
0.078–0.092 d−1. By analogy, the kd of OTC was obtained in the range of 0.131–0.144 d−1.
However, the kd of OTC was obviously higher than equal concentration of TC. For
example, the kd of TC at 0.5 mg L−1 in the presence and absence of strain ES2-1 were
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,
respectively. As the results the reflected by kinetic parameters, the degradation rate of TCs
had no correction with their initial concentrations, and that OTC was degraded faster than
3.4. Metabolites and metabolic pathways of E2 by strain ES2-1 in aqueous solution
In aqueous solution with TCs, eleven E2 metabolites belonging to three possible
degradation pathways were identified by UPLC-HRMS (Table 1); the corresponding
spectra of these metabolites were shown in Figure S1. m/z [M+H]+ found at 271.1690
(product P1, E1), m/z 287.1633 (product P2, 4-OH-E1), m/z 319.1535 (product P4), and
m/z 369.1560 (product P7) proved the existence of pathway 1, through which E2 was 14
cleaved until the complete opening of the steroidal nucleus. Besides, owing to the m/z
[M+H]+ found at 272.1635 (product N1), 304.1575 (products N3), and 308.1545 (products
N5), it was proved that there existed a pathway 2 differing from pathway 1, through which
part of phenolic ring-opening products could condense with NH3 to form the pyridine
derivatives, accompanied by decarbonylation at C4. On the contrary, decarbonylation
might not occur as the identification of products Ns1 (pyridinestrone acid) through Ns4,
with m/z values found at 300.1573, 316.1525, 332.1528, and 350.1649, respectively.
Consequently, pathway 3 was proposed, through which all metabolites performed a special
structure of pyridine carboxylic acid.
E2 ring cleavage was influenced by TCs, especially at high-level concentrations, as
reflected by the decreased abundance of ring-opening metabolites (Figure 4). More than
ten products were observed in the reaction system exposed to TC at 0.5 mg L−1 (Figure 4a),
whereas at 8 mg L−1, the abundance of metabolites was lower, but an obvious amassing of
products P2, N1, and Ns1 occurred (Figure 4b). Results illustrated that increased
concentrations of TC dramatically restrained the production of oxidative ring-opening
products but urged various pyridine derivatives to accumulate. The effect of OTC on E2
metabolism resembled that of TC but was slightly more severe. Specifically, total peak
area of the eleven products in the groups treated with 0.5 mg L−1 OTC was lower than that
of groups with 0.5 mg L−1 TC, and it was barely weakened over time (Figure 4c); When
OTC raised to 8 mg L−1, ring-opening products such as P4, N3, and Ns2–Ns4 were
scarcely detected, whereas P2 and Ns1 accumulated significantly, even when compared to
their accumulation under an equal concentration of TC (Figure 4d). These findings 15
suggested that, first, the suppression of E2 biodegradation by OTC was more difficult to
diminish with time than that of TC. Second, enhancing the concentration of TCs could
remarkably restrain E2 ring opening, but promoting the condensation of some phenol ring
opening products with NH3 in the reaction system, leading to the accumulation of pyridine
3.5. Removal of estrogens by strain ES2-1 in natural water containing TCs
To assess the removal ability of E2 by strain ES2-1 in realistic aquatic systems,
laboratory experiments based on oligotrophic environments containing TCs were designed.
Water samples were collected from three water systems flowing through Nanjing,
including the Qinhuai River, Yangtze River, and Xuanwu Lake. The basic information of
water samples was listed in Table S2. TC and OTC concentrations in the three water
samples ranged from 0.20 to 3.41 µg L−1, and from 0.72 to 3.85 µg L−1, respectively (Table
2a). Concentrations of E3, E2, and E1 in the three water samples were 0.77–7.77 µg L−1,
0.17–0.88 µg L−1, and non-detected (ND)–7.80 µg L−1, respectively. 7 d later, residual E3,
E2, and E1 concentrations in the three water samples without inoculation with strain
ES2-1 were slightly decreased, by 10.89–20.35%, 10.45–16.80%, and 13.42–19.25%,
respectively. When the water samples were treated with strain ES2-1 for 7 d, residual E3,
E2, and E1 concentrations were reduced to ND–0.61 µg L−1, ND–0.08 µg L−1, and 0–0.15
µg L−1, with removal rates of 72.38–100%, 52.10–100%, and 98.06–100%, respectively
(Table 2b). The results showed that the strain could still degrade estrogens at
environment-related concentrations in natural oligotrophic waters that might contain TCs.
For a water system potentially polluted by organic pollutants, microbial degradation
is critical in reducing the risk of such pollution. However, studies tend to focus on the
biodegradation of one type of pollutant while ignoring the possible impacts brought about
by coexisting pollutants. Estrogens and antibiotics are two groups of pollutants of concern
in large-scale farming. Therefore, this study investigated the effects of the presence of TCs
(representing antibiotics) on the biodegradation mechanisms of E2 (representing
2 mg L−1 was the minimum inhibitory concentration (MIC) of TCs on
Novosphingobium sp. ES2-1, the degrader of interest in this study (Figure S2). E2
biodegradation under the actions of low-level (≤ 2 mg L−1, represented by 0.5 mg L−1) TCs
was quite different from that of under high concentrations (≥ 2 mg L−1, represented by 8
mg L−1). Cell growth was slightly disturbed by low-level TCs but recovered in later
reaction stage, and the degradation behavior could still be described using
pseudo-first-order kinetics model. This was because, first, further reduction in low-level
TCs relieved the already low toxicity to functional strain. Second, low-level antibiotics
belonged to concentrations that lower than the MIC (sub-MIC) and could select resistance
strains. By duplications of microorganisms, a pre-existing weak resistance phenotype
could be amplified, thereby increasing the probability of bacterial survival and growth in
the presence of antibiotics (Andersson and Hughes, 2014). By contrast, bacteria exposed
to high-level antibiotic were harder to survive. Hydrolysis alone was insufficient to rapidly
reduce the toxicity of high-level TCs on microorganisms to safe levels. A slowly- or 17
non-growing state was therefore needed for bacteria to be alive with such level of
antibiotics (Allison et al., 2011). Early in the reaction stage, high-level TCs severely
suppressed the cell growth or even killed the population. A great restriction on the initial
biotransformation rate was thus formed, thus making pseudo-first-order kinetic unsuitable
for describing the biodegradation behavior under high-level TCs.
The reduction in TCs occurred simultaneously with E2 biodegradation, and its
manner of decay could also be described by pseudo-first-order kinetic model. However,
the decrease of TCs had no connection with biodegradation or photodegradation, making
hydrolysis probably the best explanation for such decline. Xuan et al. (2009) also reported
the hydrolysis of TCs in deionized water and fitted the hydrolysis curve of OTC with the
same model, and hydrolysis rate constant (kh) was obtained in the range of 0.094–0.106
d−1 at 25 °C; While the kh were within the range of 0.134–0.142 d−1 in our results but at
30 °C, the difference could be attribute to the reaction temperature. As TCs are heat-labile,
high temperatures accelerate the hydrolysis rate (Leffler and Grunwald, 1963). As
amphoteric molecules, the kh of TCs is also related to pH value of the reaction system.
TCs are easily hydrolyzed under neutral, slightly acidic, and alkaline aqueous conditions,
with aqueous half-lives ranging from 20 h to 50 d (Doi et al., 2000; Pouliquen et al., 2007;
Loftin et al., 2008; Xuan et al., 2009; Kang et al., 2010). The pH value established in this
study was close to neutral and fluctuated little during the reaction, which satisfied the
optimal requirement of TCs hydrolyzation.
The presence of TCs reduced not only the degradation efficiency but also the
abundance of ring-opening metabolites. The proposed pathways were shown in Figure 5. 18
Product P3 that probably generated via the 4,5-seco pathway (Chen et al., 2017) was
suggested to be the metabolite between P2 and P4, but the m/z evidence supporting its
generation in the presence of TCs was lacking. Product P5 and P6 were both downstream
of P4, where P5 might be produced through the 9,10-seco pathway (Horinouchi et al.,
2004; Sih et al., 1965), whereas their existences were also failed to be supported by mass
evidence when TCs were added. Similarly, the production of the proposed derivate
products N2 and N4 in the presence of TCs was not well supported. A particularly
interesting case was the accumulation of some pyridine derivatives. Ns1 was documented
in previous reports, and its formation mechanism was ascribed to the condensation of
phenol ring-cleavage products with ammonium ions (Dagley et al., 1960; Coombe et al.,
1966). The presence of TCs, especially at raised concentrations, significantly hindered the
ring-opening process but improved the contact probability between some phenolic
ring-opening products and ammonium ions, enabling E2 metabolism to linger in the stage
of condensation reaction. Moreover, the abundances of metabolites generated through
pathway 2 were obviously lower than those in pathway 3. To our knowledge, these two
pathways differ only in the loss of a CO2 molecule, which might be caused by certain
decarboxylase secreted by strain ES2-1 during the condensation reaction. It is well known
that TCs can weaken the interaction between ribosomes and tRNAs, impeding protein
synthesis (Epe et al. 1987), and the decarboxylase might be involved. This hypothesis
deserves further investigation. Remarkably, both N1 and Ns1 shared the same steroid
nucleus, yet nitrogen was introduced in such nucleus, so it remained to be seen whether
these derivatives still performed estrogenic activity. Furthermore, N1 and Ns1 might be 19
produced in any aqueous solution contained ammonium ions, the presence of antibiotics
just facilitated their accumulation. Hence, it may be an emerging factor deserved to be
considered in assessing the ecological risk.
Concentrations of natural estrogens in real water bodies are usually at ng L−1–µg L−1
levels (Furuichi et al., 2004; Laganà et al., 2004; Lei et al., 2009; Kolodziej et al., 2004;
Sarmah et al., 2006; Shappell et al., 2007; Velicu and Suri, 2009), the levels of estrogens
detected in this study generally coincided with those reported previously. Although TCs at
µg L−1 levels were detected in this study, it remained within the reasonable concentration
range. Nevertheless, such level of antibiotics might enable microbes to generate and
maintain resistance. More consideration could be given to the two-sidedness of functional
degrading bacteria developing drug resistance when assessing the ecological risks of
combined contamination. In oligotrophic water bodies containing TCs, the effective
degradation of estrogens by functional strain was primarily to blame the instability of TCs
due to hydrolysis, etc., causing the impermanence of their toxic effects on microorganisms.
Second, the oligotrophic environments starved functional microorganisms, obliging them
to utilize more organic pollutants to meet their survival needs (Marshall et al., 2000).
Third, strain ES2-1 was better adapted to low-nutrition conditions than those
microorganisms survived in nutritious circumstances, as it was isolated from a STP
containing various pollution factors.
436 437 438
5. Conclusions Levels of TCs have an important influence on the degradation of E2 by functional 20
microorganisms. The effects of TCs at sub-MIC on E2 biodegradation were quite different
from those at concentrations above MIC. The presence of TCs reduced the degradation
rate and integrity of E2 by functional bacteria through interfering with the subsequent
ring-opening process, accompanied by the condensation of various phenolic ring-opening
products with ammonium ions and finally causing E2 metabolism to linger in the
condensation stage. Notwithstanding the levels of TCs in the natural aqueous environment
were generally low, which was insufficient to cause significant biodegradation interference,
water systems that receive high-level antibiotics might still face such uncertain risks. More
studies can focus on assessing the estrogenic activity of those derivatives, as well as
shielding functional microorganisms from the injury of TCs during biodegradation.
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
Declaration of Interest Statement
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
This work was financially supported by the National Natural Science Foundation of
China (41977121, 41771523), and the National Key Research and Development Program
of China (2016YFD0800203).
Allison, K.R., Brynildsen, M.P., Collins, J.J., 2011. Metabolite-enabled eradication of bacterial
468 469 470 471 472 473 474 475 476
persisters by aminoglycosides. Nature 473, 216. Andersson, D.I., Hughes, D., 2014. Microbiological effects of sublethal levels of antibiotics. Nat. Rev. Microbiol. 12, 465–478. Arnold, K.E., Brown, A.R., Ankley, G.T., Sumpter, J.P., 2014. Medicating the environment: assessing risks of pharmaceuticals to wildlife and ecosystems. Phil. Trans. R. Soc. B 369, 20130569. Adeel, M., Song, X., Wang, Y., Francis, D., Yang, Y., 2017. Environmental impact of estrogens on human, animal and plant life: a critical review. Environ. Int. 99, 107–119. Beck, I.C., Bruhn, R., Gandrass, J., 2006. Analysis of estrogenic activity in coastal surface waters of the Baltic Sea using the yeast estrogen screen. Chemosphere 63, 1870–1878.
Bradley, P.M., Barber, L.B., Chapelle, F.H., Gray, J.L., Kolpin, D.W., McMahon, P.B., 2009.
Biodegradation of 17β-estradiol, estrone and testosterone in stream sediments. Environ. Sci.
Technol. 43, 1902–1910.
480 481 482
Coombe, R.G., Tsong, Y.Y., Hamilton, P.B., Sih, C.J., 1966. Mechanisms of steroid oxidation by microorganisms X. Oxidative cleavage of estrone. J. Biol. Chem. 241, 1587–1595. Chen, Y.L., Yu, C.P., Lee, T.H., Goh, K.S., Chu, K.H., Wang, P.H., Ismail, W., Shih, C.J., Chiang, Y.R., 22
2017. Biochemical mechanisms and catabolic enzymes involved in bacterial estrogen degradation
pathways. Cell Chem. Biol. 24, 1–13.
Dagley, S., Evans, W.C. Ribbons, D.W., 1960. New pathways in the oxidative metabolism of aromatic compounds by micro-organisms. Nature 188, 560–566.
Doi, A.M., Stoskopf, M.K., 2000. The kinetics of oxytetracycline degradation in deionized water under
varying temperature, pH, light, substrate, and organic matter. J. Aquat. Anim. Health 12, 246–253.
Epe, B., Woolley, P., Hornig, H., 1987. Competition between tetracycline and tRNA at both P and A sites of the ribosome of Escherichia coli. FEBS Lett. 213, 443–447.
EE Commission, 2013. Directive 2013/39/EU of the European Parliament and of the Council of 12
August 2013 amending Directives 2000/60/EC and 2008/105/EC as regards priority substances in
the field of water policy. Off. J. Eur Union L226, 1–17.
Fujii, K., Kikuchi, S., Satomi, M., Ushio-Sata, N., Morita, N., 2002. Degradation of 17β-estradiol by a
gram-negative bacterium isolated from activated sludge in a sewage treatment plant in Tokyo,
Japan. Appl. Environ. Microbiol. 68, 2057–2060.
Furuichi, T., Kannan, K., Giesy, J.P., Masunaga, S., 2004. Contribution of known endocrine disrupting
substances to the estrogenic activity in Tama River water samples from Japan using instrumental
analysis and in vitro reporter gene assay. Water Res. 38, 4491–501.
Gracia, T., Hilscherova, K., Jones, P.D., Newsted, J.L., Higley, E.B., Zhang, X., Hecker, M., Murphy,
M.B., Yu, R.M.K., Lam, P.K.S., Wu, R.S.S., Giesy, J.P., 2007. Modulation of steroidogenic gene
expression and hormone production of H295R cells by pharmaceuticals and other environmentally
active compounds. Toxicol. Appl. Pharmacol. 255, 142–153.
Holm, J.V., Rügge, K., Bjerg, P.L., Christensen, T.H., 1995. Occurrence and distribution of 23
pharmaceutical organic compounds in the groundwater downgradient of a landfill (Grindsted,
Denmark). Environ. Sci. Technol. 29, 1415–1420
Horinouchi, M., Hayashi, T., Kudo, T., 2004. The genes encoding the hydroxylase of
3-hydroxy-9,10-secoandrosta-1,3,5(10)-triene-9,17-dione in steroid degradation in Comamonas
testosteroni TA441. J. Steroid Biochem. Mol. Biol. 92, 143–154.
Hashimoto, T., Onda, K., Morita, T., Luxmy, B.S., Tada, K., Miya, A., Murakami, T., 2010.
Contribution of the estrogen-degrading bacterium Novosphingobium sp. strain JEM-1 to estrogen
removal in wastewater treatment. J. Environ. Eng. 136, 890–896.
He, Y., Wang, T., Sun, F., Wang, L., Ji, R., 2019. Effects of veterinary antibiotics on the fate and persistence of 17β-estradiol in swine manure. J. Hazard. Mater. 375, 198–205.
Jenkins, M.B., Endale, D.M., Schomberg, H.H., Sharpe, R.R., 2006. Fecal bacteria and sex hormones
in soil and runoff from cropped watersheds amended with poultry litter. Sci. Total Environ. 358,
Jiang, Y., Li, M., Guo, C., An, D., Xu, J., Zhang, Y., Xi, B., 2014. Distribution and ecological risk of
antibiotics in a typical effluent–receiving river (Wangyang River) in north China. Chemosphere
Kolodziej, E.P., Harter, T., Sedlak, D.L., 2004. Dairy wastewater, aquaculture, and spawning fish as
sources of steroid hormones in the aquatic environment. Environ. Sci. Technol. 38, 6377–6384.
Karthikeyan, K.G., Meyer, M.T., 2006. Occurrence of antibiotics in wastewater treatment facilities in
Wisconsin, USA. Sci. Total Environ. 361, 196–207.
Khanal, S.K., Xie, B., Thompson, M.L., Sung, S., Ong, S.K., Van Leeuwen, J., 2006. Fate, transport,
and biodegradation of natural estrogens in the environment and engineered systems. Environ. Sci. 24
Technol. 40, 6537–6546.
Kjær, J., Olsen, P., Bach, K., Barlebo, H.C., Ingerslev, F., Hansen, M., Sørensen, B.H., 2007. Leaching
of estrogenic hormones from manure-treated structured soils. Environ. Sci. Technol. 41, 3911–
Kolodziej, E.P., Sedlak, D.L., 2007. Rangeland grazing as a source of steroid hormones to surface waters. Environ. Sci. Technol. 41, 3514–3520.
Kurisu, F., Ogura, M., Saitoh, S., Yamazoe, A., Yagi, O., 2010. Degradation of natural estrogen and
identification of the metabolites produced by soil isolates of Rhodococcus sp. and Sphingomonas
sp. J. Biosci. Bioeng. 109, 576–582.
Kang, H.J., Lim, M.Y., Kwon, J.H., 2012. Effects of adsorption onto silica sand particles on the hydrolysis of tetracycline antibiotics. J. Environ. Monitor. 14, 1853–1859.
Laganà, A., Bacaloni, A., De Leva, I., Faberi, A., Fago, G., Marino, A., 2004. Analytical methodologies
for determining the occurrence of endocrine disrupting chemicals in sewage treatment plants and
natural waters. Anal. Chim. Acta 501, 79–88.
541 542 543 544 545 546
Loftin, K.A., Adams, C.D., Meyer, M.T., Surampalli, R., 2008. Effects of ionic strength, temperature, and pH on degradation of selected antibiotics. J. Environ. Qual. 37, 378–386. Lei B, Huang S, Zhou Y, Wang D, Wang Z., 2009 Levels of six estrogens in water and sediment from three rivers in Tianjin area, China. Chemosphere 76, 36–42. Leffler, J.E., Grunwald E., 1963. Rates and equilibria of organic reactions: as treated by statistical, thermodynamic and extrathermodynamic methods.
Leng, Y.F., Bao, J.G., Chang, G.F., Zheng, H., Li, X.X., Du, J.K., Snow, D., Li, X., 2016.
Biotransformation of tetracycline by a novel bacterial strain Stenotrophomonas maltophilia DT1. J. 25
Hazard. Mater. 318, 125–133.
Liu, Y., He, X., Duan, X., Fu, Y., Fatta-Kassinos, D., Dionysiou, D.D., 2016. Significant role of UV and
carbonate radical on the degradation of oxytetracycline in UV-AOPs: kinetics and mechanism.
Water Res. 95, 195-204.
Li, S.Y., Liu, J., Sun, M.X., Ling, W.T., Zhu, X.Z., 2017. Isolation, characterization, and degradation
performance of the 17β-estradiol-degrading bacterium Novosphingobium sp. E2S. Int. J. Env. Res.
Pub. He. 14, 115.
Liu, J., Li, S.Y., Li, X., Gao, Y.Z., Ling, W.T., 2018. Removal of estrone, 17β-estradiol, and estriol from
sewage and cow dung by immobilized Novosphingobium sp. ARI-1. Environ. Technol. 39, 2423–
559 560 561 562
Marshall, S.J., House, W.A., White, G.F., 2000. Role of natural organic matter in accelerating bacterial biodegradation of sodium dodecyl sulfate in rivers. Environ. Sci. Technol. 34, 2237–2242. Matsui, Y., Ozu, T., Inoue, T., Matsushita, T., 2008. Occurrence of a veterinary antibiotic in streams in a small catchment area with livestock farms. Desalination 226, 215–221.
Pouliquen, H., Delépée, R., Larhantec-Verdier, M., Morvan, M.L., Le Bris, H., 2007. Comparative
hydrolysis and photolysis of four antibacterial agents (oxytetracycline oxolinic acid, flumequine
and florfenicol) in deionised water, freshwater and seawater under abiotic conditions. Aquaculture
Pal, A., Gin, K.Y.H., Lin, A.Y.C., Reinhard, M., 2010. Impacts of emerging organic contaminants on
freshwater resources: review of recent occurrences, sources, fate and effects. Sci. Total Environ.
Qi, W., Long, J., Feng, C., Feng, Y., Cheng, D., Liu, Y., Xue, J., Li, Z. 2019. Fe3+ enhanced degradation 26
of oxytetracycline in water by Pseudomonas. Water Res. 160, 361–370.
Rocha, M. J., Ribeiro, M., Ribeiro, C., Couto, C., Cruzeiro, C., Rocha, E., 2012. Endocrine disruptors
in the Leça River and nearby Porto Coast (NW Portugal): presence of estrogenic compounds and
hypoxic conditions. Toxicol. Environ. Chem. 94, 262–274.
Ray, P., Zhao, Z., Knowlton, K. F., 2013. Emerging contaminants in livestock manure: hormones, antibiotics and antibiotic resistance genes. Sustain. Anim. Agric. 268–283.
Rose, E., Paczolt, K.A., Jones, A.G., 2013. The effects of synthetic estrogen exposure on premating and
postmating episodes of selection in sex-role-reversed Gulf pipefish. Evol. Appl. 6, 1160–1170.
Sih, C.J., Wang, K.C., Gibson, D.T., Whitlock Jr, H.W., 1965. On the mechanism of ring A cleavage in
the degradation of 9, 10-seco steroids by microorganisms. J. Am. Chem. Soc. 87, 1386–1387.
Sarmah, A.K., Northcott, G.L., Leusch, F.D., Tremblay, L.A., 2006. A survey of endocrine disrupting
chemicals (EDCs) in municipal sewage and animal waste effluents in the Waikato region of New
Zealand. Sci. Total. Environ. 355, 135–144.
Swartz, C. H., Reddy, S., Benotti, M. J., Yin, H. F., Barber, L. B., Brownawell, B. J., Rudel, R.A., 2006.
Steroid estrogens, nonylphenol ethoxylate metabolites, and other wastewater contaminants in
groundwater affected by a residential septic system on Cape Cod, MA. Environ. Sci. Technol. 40,
Shappell, N.W., Billey, L.O., Forbes, D., Matheny, T.A., Poach, M.E., Reddy, G.B., 2007. Estrogenic
activity and steroid hormones in swine wastewater through a lagoon constructed-wetland system.
Environ. Sci. Technol. 41, 444–450.
Tetreault, G.R., Bennett, C.J., Shires, K., Knight, B., Servos, M.R., McMaster, M.E., 2011. Intersex and
reproductive impairment of wild fish exposed to multiple municipal wastewater discharges. Aquat. 27
593 594 595 596 597
Toxicol. 104, 278–290. US-EPA,
https://www.epa.gov/ccl/chemical-contaminants-ccl-4 Velicu, M., Suri, R., 2009. Presence of steroid hormones and antibiotics in surface water of agricultural, suburban and mixed-use areas. Environ. Monit. Assess. 154, 349–359.
Wicks, C., Kelley, C., Peterson, E., 2004. Estrogen in a karstic aquifer. Groundwater, 42, 384–389.
Wocławek-Potocka, I., Mannelli, C., Boruszewska, D., Kowalczyk-Zieba, I., Waśniewski, T.,
Skarżyński, D.J., 2013. Diverse effects of phytoestrogens on the reproductive performance: cow as
a model. Int. J. Endocrinol.
Xu, W.H., Zhang, G., Li, X.D., Zou, S.C., Li, P., Hu, Z.H., Li, J., 2007. Occurrence and elimination of
antibiotics at four sewage treatment plants in the Pearl River Delta (PRD), South China. Water Res.
605 606 607 608 609 610
Xuan, R.C., Arisi, L., Wang, Q.Q., Yates, Scott R, Biswas, Keka C., 2009. Hydrolysis and photolysis of oxytetracycline in aqueous solution. J. Environ. Sci. Heal. B 45, 73–81. Ying, G.G., Kookana, R.S., Ru, Y.J., 2002. Occurrence and fate of hormone steroids in the environment. Environ. Int. 28, 545–551. Yu, C.P., Roh, H., Chu, K.H., 2007. 17β-estradiol-degrading bacteria isolated from activated sludge. Environ. Sci. Technol. 41, 486–492.
Yang, C., Yu, W., Bi, Y., Long, F., Li, Y., Wei, D., Hao. X., Suit, J., Zhao, Y., Huang, F., 2018. Increased
oestradiol in hepatitis E virus infected pregnant women promotes viral replication. J. Viral Hepat.
Zhao, L., Dong, Y.H., Wang, H., 2010. Residues of veterinary antibiotics in manures from feedlot 28
livestock in eight provinces of China. Sci. Total Environ. 408, 1069–1075.
Legends of figures and tables
Figure 1. E2 Biodegradation efficiency of strain ES2-1 in aqueous solution with TCs at
various concentrations. (a) Influence of different TC concentrations on E2
biodegradation efficiency; (b) Residual TC and cell counts of strain ES2-1 in the
degradation system after 7 d; (c) Influence of different OTC concentrations on E2
biodegradation efficiency; (d) Residual OTC and biomass of strain ES2-1. * p < 0.05,
** p < 0.01, Duncan’s test, control (without ES2-1) versus experimental groups (with
ES2-1) at different concentrations of TCs.
Figure 2. E2 biodegradation kinetics in the presence of low- or high-level TCs represented
by 0.5 and 8 mg L−1 TCs, respectively. (a) E2 biodegradation kinetics without TCs;
E2 biodegradation kinetics under the action of low-level TC (b) or OTC (c); E2
biodegradation kinetics under the action of high-level TC (d) or OTC (e); (f) E2
biodegradation kinetic parameters obtained by fitting with the pseudo-first-order
kinetic model. * p < 0.05, ** p < 0.01, Duncan’s test, groups with 0.5 mg L−1 TCs
versus groups with 8 mg L−1 TCs.
Figure 3. Degradation kinetics of TCs during the biodegradation of E2. Degradation
kinetics curves of 0.5 (a) and 8 (b) mg L−1 TC; Degradation kinetics curves of 0.5 (c)
and 8 (d) mg L−1 OTC; (e) Degradation kinetics parameters of TCs with different
initial concentrations after fitting with the pseudo-first-order dynamic model. * p <
0.05, ** p < 0.01, Duncan’s test, groups treated by TC versus groups treated by OTC.
Figure 4. Abundances of E2 metabolites in the presence of TCs. Products P1, P2, P4, and
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.
Table 1. UPLC-HRMS data of E2 metabolites.
Table 2. Removal of estrogens from natural water samples containing TCs by strain ES2-1.
(a) Initial and residual (after 7 d) concentrations of TCs in three water samples; (b)
Initial and residual (after 7 d) concentrations of E3, E2, E1 in different inoculation
Table 1 Chemical
Yu et al., 2007
Chen et al., 2017
Coombe et al., 1966
Predicted chemical structure
(E1) P2 (4-OH-E1)
Chen et al., 2017 Coombe et al., 1966
Sih et al., 1965 O HOOC
This study O HOOC
(a) TCs in natural water Initial concentrations (µg L−1)
Residual concentrations (µg L−1)
0.30 ± 0.1
0.20 ± 0.01
3.41 ± 0.35
0.20 ± 0.07
1.37 ± 0.40
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 (%)
2.20 ± 0.33
0.17 ± 0.02
1.75 ± 0.13
7.77 ± 0.25
0.88 ± 0.05
7.80 ± 0.87
6.42 ± 0.26
0.77 ± 0.15
0.45 ± 0.08
0.31 ± 0.06
0.68 ± 0.09
2.20 ± 0.33
0.17 ± 0.02
0.61 ± 0.05
7.77 ± 0.25
0.88 ± 0.05
7.80 ± 0.87
0.77 ± 0.15
0.45 ± 0.08
0.31 ± 0.06
0.1 5 ± 0.03
0.78 ± 0.05
6.75 ± 1.17
0.38 ± 0.03
0.25 ± 0.02
0.08 ± 0.04
0.01 ± 0.00
0.15 ± 0.03
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.