Occurrence and Ecotoxicological Effects of Free, Conjugated, and

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Occurrence and ecotoxicological effects of free, conjugated and halogenated steroids of 17#-hydroxypregnanolone and pregnanediol in Swiss wastewater and surface water Kun Zhang, Yanbin Zhao, and Karl Fent Environ. Sci. Technol., Just Accepted Manuscript • Publication Date (Web): 09 May 2017 Downloaded from http://pubs.acs.org on May 9, 2017

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Environmental Science & Technology

1

Occurrence and ecotoxicological effects of free, conjugated and halogenated

2

steroids including 17α α-hydroxypregnanolone and pregnanediol in Swiss

3

wastewater and surface water

4 5 6

Kun Zhang, 1Yanbin Zhao, 1 Karl Fent, 1, 2*

7

1

8

Gründenstrasse 40, CH–4132 Muttenz, Switzerland

9

2

University of Applied Sciences and Arts Northwestern Switzerland, School of Life Sciences,

Swiss Federal Institute of Technology Zürich (ETH Zürich), Department of Environmental Sys-

10

tem Sciences, Institute of Biogeochemistry and Pollution Dynamics, CH-8092 Zürich, Swit-

11

zerland

12 13 14 15

*Corresponding Author

16

Phone: +41 61 467 4571; fax: +41 61 467 47 84; e-mail: [email protected];

17

ACS Paragon Plus Environment

1


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ABSTRACT

19

Apart from estrogens, the occurrence and ecotoxicity of steroids in the aquatic environment is

20

poorly known. Here we analyzed 33 steroids, including estrogens, androgens, progestins, and

21

glucocorticoids in hospital wastewaters, river water and in municipal wastewater treatment plant

22

(WTP) influents and effluents at different sites in Switzerland. In addition, wastewater from

23

different treatment steps of two WTPs with advanced treatment, such as ozonation or pulverized

24

activated carbon, were analyzed to study the steroid’s behavior during treatment. Considerable

25

levels of different steroids occurred in hospital and raw municipal wastewater but they were low

26

(lower than 1 ng/L) or below detection level in effluents of WTPs and river water. In WTP

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influents estrogens (estrone, 17β-estradiol, estriol), androgens (androstenedione, androsterone,

28

trans-androsterone, testosterone), progestins and metabolites (medroxyprogesterone acetate,

29

megestrol

30

hydroxyprogesterone, and 21α-hydroxyprogesterone) were detected and removed effectively

31

during biological treatment. Ozonation further removed the steroids. Exposure of zebrafish

32

embryos demonstrated negligible effects of pregnanediol and 17α-hydroxypregnenolone, while

33

mixtures that mimic wastewater and river water composition affected embryo development and

34

led to alteration of steroidogenesis gene transcripts at ng/L concentrations. Although steroid

35

concentrations are low in Swiss rivers, the possibility of additive effects may be of concern.

acetate,

mifepristone,

pregnanediol,

17α-hydroxypregnanolone,

36


2

17α-

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TOC

WTP er2b cyp19a er pr vtg1 cyp11b2 ar hsd11b2 cyp17

Effects in fish

37 38 39 40


3


41 42 43

INTRODUCTION

44

endocrine disrupting activities which may cause adverse reproductive effects, feminization and

45

intersex in fish even at ng/L concentrations.1-5 Previous studies focused largely on estrogens and

46

also androgens, but only recently the occurrence and behavior of other steroids such as

47

progestins were studied despite the fact that the consumption of progestins for medical use is

48

larger than that of estrogens and androgens.2, 6-14 The concentrations of some progestins such as

49

progesterone (P4), medroxyprogesterone acetate (MPA) and norethindrone (NTD) have been

50

identified and found at up to 35 ng/L, 18 ng/L and 16 ng/L, respectively, in an urban river of

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Beijing, China. Their concentrations in the influent of a Beijing wastewater treatment plant

52

(WTP) were detected up to 108 ± 89 ng/L, 58 ± 17 ng/L and 12 ± 0.6 ng/L, respectively, and they

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were generally higher than those of estrogens.11, 15 The mean concentrations in surface water

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collected from the Rhône Alpes region in France were 1.6 ng/L for P4 to 2 ng/L for NTD.9 In

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WTP influents and effluents in Belgium, the concentration of P4 was up to 33 and 2.5 ng/L,

56

respectively.16 Additional data are reported from other countries, including the United States, and

57

where present, concentrations were in the ng/L, but higher in runoffs of animal production

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farms.2 Removal of progestins was 73 to 96% in several Chinese WTPs.15 However, there is very

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little information about the occurrence and environmental behavior of progestins in Switzerland.

60

One measurement showed that P4, 21α-hydroxyprogesterone (21α-HPA), and mifepristone

61

(MFST) occurred at ng/L concentrations in a river, hospital wastewater and WTP.10

62

Consequently, the lack of information on the occurrence and behavior of steroids, including

63

progestins, metabolites and androgens in different environmental media is of concern.

Natural and synthetic steroid hormones are of environmental concern due to their

64

Steroids and their metabolites (including sulphate and glucuronide conjugates) originate

65

from excretion via urine and feces and represent the main sources in wastewater.2 Steroids could


4

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also be excreted in urine as metabolites. Pregnanediol (PD) and 17α-hydroxy pregnanolone

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(17OH-∆5P) are two major progestin metabolites, which could be excreted via urine.17

68

Concentrations of PD in urine could be higher than its precursor P4 and its concentration in

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influents of WTPs in France was predicted to be about 340 ng/L.18 We developed a trace

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analytical method to determine 17OH-∆5P, PD and other 31 steroids including estrogens,

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androgens, progestins and glucocorticoids and their metabolites, as well as 17OH-∆5P and PD,

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which were detected in WTP influents at concentrations up to 400 ng/L.19 However, their fate,

73

behavior in the treatment process and ecotoxicological consequences has not yet been evaluated.

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The aim of the present study was to determine an unprecedented series of 33 different

75

natural and synthetic steroid hormones and their metabolites in typical Swiss rivers to assess

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their occurrence in surface water. Hospital wastewater, influent and effluent samples of different

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WTPs were also studied as potential input sources. We also studied WTPs with different

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treatment technologies to obtain information about the removal efficiency of conventional and

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advanced treatment technologies. In addition, the development of zebrafish embryos, behavioral

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and transcriptional responses of target genes were investigated upon exposure of embryos to

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17OH-∆5P, PD and two mixtures representing WTP and river water concentrations to better

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understand potential ecotoxicological implications of these environmental steroids.

83 84

MATERIALS AND METHODS

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Chemicals. High purity authentic standards of 35 steroids, androstenedione (ADD),

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androsterone (ADR), trans-androsterone (TADR), testosterone (TTR), Spironolactone (SPL),

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estrone (E1), 17β-estradiol (E2), estriol (E3), ethinyl estradiol (EE2), corticosterone (COR),

88

fludrocortisone

acetate

(FLUA),

17α-hydroxyprogesterone


5

(17α-HPA),

21α-


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89

hydroxyprogesterone (21α-HPA), chlormadinone (CM), chlormadinone acetate (CMA),

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cyproterone acetate (CPA), dydrogesterone (DDG), dienogest (DNG), drospirenone (DRS),

91

gestodene (GES), medroxyprogesterone (MP), medroxyprogesterone acetate (MPA), megestrol

92

acetate (MTA), 19-norethindrone (NTD), progesterone (P4), mifepristone (MFST), 17α-

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hydroxyprogesterone-d8 (17α-HPA-d8), 17β-estradiol-d5 (E2-d5), progesterone-d9 (P4-d9) were

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purchased

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(monoBrEE2), 2-bromo-17β-estradiol (monoBrE2), 17β-estradiol-3-disulfate (E2-3S), 17α-

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ethynylestradiol-3-glucuronide

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hydroxypregnanolone (17OH-∆5P), pregnanediol (PD) were purchased from Steraloids

98

(Newport, USA). 17β-Estradiol-3-Sulfate-d4 (E2-S-d4) was purchased from EQ Laboratories

99

GmbH (Augsburg, Germany). Standards were dissolved in methanol at concentration of 5

100

mg/mL (stock solution) Working solutions were prepared fresh before each analytical run by

101

diluting stock solution in methanol to concentration of 10 ng/mL. All stock and working

102

solutions were stored at -20 °C in the dark.

from

Sigma-Aldrich

(Buchs,

(EE2-3G),

Switzerland).

2-bromo-17α-ethynylestradiol

17α-ethynylestradiol-3-sulfate

(EE2-3S),

17α-

103

Sample collection. Grab samples were collected at different times in 2015 and 2016 in

104

the wastewater outflow of the general hospital of the city of Wetzikon, and also in wastewater of

105

three hospitals of other cities. Sampling locations and sampling times are listed in Table S1

106

(Supplementary information). Time-proportional 24 h influent and effluent samples were

107

collected from nine different WTPs at different times in 2015 and 2016. Time-proportional 24 h

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samples from the influent, before advanced treatment and effluent were collected at different

109

times in 2015 and 2016 in two WTPs. WTP Neugut, Dübendorf, was equipped with ozonation,

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and WTP Herisau with pulverized activated carbon treatment. Water samples in 1 L clean glass

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bottles were transported to the laboratory and their pH was adjusted to 3 with acetic acid. The


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samples were stored in the dark at 4°C, and processed within 24 h. Flow-proportional half-year

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mixtures of influent and effluent samples were collected in the communal WTP of the city of

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Basel (WTP Prorheno AG) in 2014. These half-year mixtures samples were made by daily

115

samples were collected with outflow-proportional samplers and frozen in a bottle for monthly

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collection sample at -20°C. Six monthly collection samples from January to June and from July

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to December 2014 were then mixed as a half-year composite sample.

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Grab river water samples of 1 L were taken monthly from March to June 2015, and

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February to April 2016 from the rivers Birs (canton Basel-Land), Aabach, Wildbach and

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Schwarz (canton Zürich). Weekly flow-proportional water samples were taken from the river

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Rhine (canton Basel-Stadt). Sampling locations in the canton Basel and Zürich are depicted in

122

Figure 1.

123

Sample Extraction and Instrument Analysis. Sample extraction and instrumental 19

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analysis were carried out following analytical method established previously

125

samples (100 mL for influent and hospital waste water, 500 mL for effluent and 1L for river

126

water) were extracted with HLB cartridges and then rinsed with two mixtures of MeOH and

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Milli-Q water. After dried with nitrogen, HLB cartridges were connected to NH2 cartridges and

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eluted sequentially with EtOAc, ACN/EtOAc (1:1, v/v) and 2% ammonia in MeOH. The extracts

129

were analyzed with HPLC-MS/MS. More details can be found in the Supplementary

130

Information. Method detection limits (MDLs) and method quantification limits (MQLs) were

131

listed in Table S3, which were determined by spiking standards to different matrixes, and

132

estimated based on the water sample enrichment factor (250 for influent, 1250 for effluent and

133

2500 for river water samples) and the signal to noise (S/N) ratio set as 3 and 10, respectively. In

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the samples of 2015, a total of 20 steroids and metabolites were analyzed, while in the samples


7

. Briefly, water


135 136

of 2016, a total of 33 steroids were determined. Zebrafish Embryo Exposure. Exposure of zebrafish embryos was performed as 20, 21

137

previously

with slight modifications. In brief, at 2−3 h post fertilization (hpf), 50 blastula-

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stage embryos per replicate (four replicates for each treatment group) were randomly placed in

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150 mL covered glass beakers containing 100 mL of reconstituted fish water (salts: CaCl2×2H2O

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147.0 g/L, KCl 2.9 g/L, MgSO4×7H2O 61.6g/L, NaHCO3 32.4 g/L) at 27 ± 1 °C, with a

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conductivity of 470–480 µS/cm. The experimental setup consisted of the following exposure

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treatments: solvent control (0.01% DMSO), 300 ng/L 17OH-∆5P, 1000 ng/L PD, an artificial

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mixture representing wastewater concentrations (Mix-W containing 20 ng/L P4, 50 ng/L TTR,

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50 ng/L PRE, 100 ng/L ADD, 600 ng/L ADR, 10 ng/L E2, 25 ng/L E1, 300 ng/L PD and 100

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ng/L 17OH-∆5P) and a mixture representing river water (Mix-R with 1 ng/L P4, 1 ng/L ADD

146

and 1 ng/L 17OH-∆5P). The 17OH-∆5P and PD concentrations were selected based on their

147

highest concentrations in WTP influents, while the concentrations of artificial mixtures, were

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selected to reflect environmentally realistic concentrations in wastewater influents and rivers,

149

respectively, as determined in the present study. A 24 h semi-static procedure was applied for the

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exposure. Every 24 h, lethal and sublethal effects were evaluated, and dead embryos were

151

removed. Water was completely changed every 24 hours with the new reconstituted fish water

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with appropriate concentrations of steroids. At 24 hpf the contraction rate, at 36 hpf heartbeat

153

and at 48 hpf the hatching success were measured. After termination of the experiment, 15

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eleuthero-embryos from each treatment were pooled and stored in RNAlater (Qiagen, Basel,

155

Switzerland) for further molecular analysis.

156

Analysis of Swimming Behavior. Swimming behavior of 72 hpf zebrafish eleuthero-

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embryos was performed as described previously.20 In brief, the ViewPoint behavior recording


8

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158

system, together with an associated movement tracking and analysis software (Zebrabox,

159

ViewPoint Life Sciences, France) were employed. For each treatment, 8 eleuthero-embryos were

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selected randomly (two eleuthero-embryos from each of the four replicates), and gently placed

161

into a 48-well cell culture plate. After an initial 10 minutes acclimation period, the swimming

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behavior activity was recorded for the next 10 minutes. Light intensity was controlled and was

163

constant during the behavior measurements.

164

RNA Isolation and qRT-PCR Analysis. RNA isolation, first strand cDNA synthesis

165

and the qRT-PCR analysis were performed according to methods described previously.20, 21 In

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brief, total RNA was extracted from a pool of 15 zebrafish eleuthero-embryos (48 hpf) by use of

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RNeasy Mini Kit (Qiagen, Basel, Switzerland). The samples were then treated with RNase-free

168

DNase (Qiagen, Basel, Switzerland) to purify the RNA from DNA contamination. RNA

169

concentrations and qualities were analyzed and then stored at −80 °C for RT-qPCR analysis.

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β-actin was selected as the reference gene for normalization due to its high gene

171

expression stability in zebrafish embryos in response to steroids.20,

21

172

employed for primers design. The intron/exon boundary spanning primers were selected to

173

minimize DNA contamination. Melting curves were analyzed to ensure that only a single product

174

was amplified. Threshold cycle (CT) values were recorded in the linear phase of amplification

175

and the data were analyzed using the delta−delta CT method of relative quantification.20,

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Further details are found in the SI.

Primer-BLAST was

22

177 178

Data Analysis and Statistics. In zebrafish experiments, the significance of differences

179

between the solvent control and each treatment in embryo development parameters and transcript

180

levels were analyzed by one-way analysis of variance (ANOVA) followed by Tukey post-hoc


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181

test (95% confidence interval). Before running the ANOVA, data was tested for the normality

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and homogeneity of variances using Shapiro-Wilk and Levene’s tests, respectively. Log-

183

transformation was performed when necessary. Hierarchical clustering map was constructed by

184

use of MultiExperimental Viewer v4.9 (Dana-Farber Cancer Institute, Boston). Results are given

185

as mean ± standard deviation (S.D.). Differences were considered as significant when p < 0.05.

186 187

RESULTS AND DISCUSSION

188

Occurrences of Steroid Hormones in Hospital Wastewater.

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In hospital Wetzikon wastewater, which was sampled from 9:00 a.m. to 5:00 p.m., 14

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steroids, including four estrogens (E1, E2, E3 and EE2), four androgens (ADD, ADR, TADR and

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TTR), one glucocorticoid (COR), six progestins and metabolites (P4, 17α-HPA, 21α-HPA, DRS,

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PD and 17OH-∆5P) were detected (Figure 1, Table S3). Among them, ADD, TTR, E1, E3 and

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P4 were always detected, followed by 17OH-∆5P, 17α-HPA, PD, and E2 with a detection

194

frequency of 89%, 78%, 67% and 56%, respectively.

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In the nine samples collected on October 28 2015, the maximum total steroid

196

concentration was 2200 ng/L (Table S3). The concentration ranges of ADD, TTR, E1, E3 and P4

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were 5.1-130 ng/L, 1.0-39 ng/L, 5.4-25 ng/L, 15-530 ng/L and 1.6-260 ng/L, respectively. The

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predominant steroids were ADD, TTR, E1, E3, P4, 17OH-∆5P and PD with an average

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contribution of 8.1%, 1.7%, 3.9%, 17%, 4.1%, 16% and 18% although their concentration varied

200

in time. E1 and E3 were the predominant estrogens, which is consistent with their occurrence in

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urine.23, 24 EE2 was detected only in two samples and DRS was measured in only one sample (13

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ng/L). Although the detection frequency of TADR and ADR were lower than that of ADD and

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TTR, its concentration was higher at up to 750 ng/L, which is consistent with their profile in the


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urine.23, 25 The low detection frequency may be due to the elevated detection limit of ADR and

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TADR (40 ng/L for ADR and 40 ng/L for TADR).

206

The highest concentrations of these steroids were comparable to those detected in

207

hospital wastewater in Belgium,16 and in Texas, USA.26 Our study reported the occurrence and

208

considerable concentrations of PD and 17OH-∆5P for the first time. These concentrations could

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be more than 50-times higher than their corresponding parent steroids P4 and 17-HPA, which is

210

consistent with the reported occurrence of PD in the urine.23

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As shown in Figure 2, total concentrations of steroids over one day show high variations

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from 28 to 2200 ng/L, with maximal concentration at 17:00 and minimal concentration at 15:00.

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For TTR, E3 and P4, higher concentrations were found in samples at 16:00 and 17:00. For ADD,

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E1 and 17OH-∆5P, higher concentration were found both in early morning (9:00) and late

215

afternoon (16:00 and 17:00).

216

In the six samples collected in 2015 from hospital Wetzikon, Kantonsspital Liestal,

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Kantonsspital Aarau and Kantonsspital Baselland, TTR, COR and P4 were detected at

218

concentrations of up to 91 ng/L, 60 ng/L and 49 ng/L, respectively. P4 occurred in all samples.

219

Generally, steroid concentrations varied considerably over time and between hospitals.

220 221

Occurrence of Steroid Hormones in municipal wastewater. Steroid concentrations were

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analyzed in 24 h composite influent and effluent samples of in six WTPs in the Basel area

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(Birsfeld, Birsig, Sissach, Füllinsdorf, Niederdorf and Bubendorf) in 2016. All WTPs were

224

operated with conventional activated sludge treatment. Of the 33 steroids, 14 were detected in

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the influent of WTP Birsfeld, Birsig, Sissach, Füllinsdorf, Niederdorf and Bubendorf, including

226

three estrogens (E1, E2 and E3), four androgens (ADD, ADR, TADR and TTR), five progestins


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(P4, 17α-HPA, 21α-HPA, MPA, MTA), and two progestin metabolites (PD and 17OH-∆5P), but

228

none of the three glucocorticoids were above the detection limit (Fig. 3, Table S4). ADD, TTR,

229

E1, E2, E3 and P4 were most frequently detected with a frequency over 75% in WTP influent

230

samples, followed by ADR, 17OH-∆5P, TADR, 17α-HPA and PD with a detection frequency of

231

50%, 33%, 25%, 16% and 16%, respectively. The similar pattern of predominant steroids in

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hospital wastewater and municipal wastewater may be related to the fact that first, naturally

233

excreted steroids occurred, and second, that most of the steroid consumption was for hormone

234

replacement therapy or contraceptives. Thus, this excretion pattern resembles that of the general

235

population taking the same steroids as in the hospitals. Pauwels et al. (2008) compared the

236

difference between the steroid hormone concentrations in domestic and hospital wastewaters and

237

no significant difference between household and hospital wastewater occurred. 16

238

Of three estrogens analyzed, E3 showed the highest mean concentration and occurred in

239

almost all of the influent samples, followed by E1, and E2. Compared to other steroids, the

240

concentrations of E1 and E3 were relatively constant in the WTP influents with relative standard

241

deviation (RSD) of 51% and 68%, respectively. Their concentrations were comparable to those

242

observed for instance in Rome, Italy 6 but higher than those in Beijing, China. 15 The dominance

243

of E3 and E1 was expected due to their abundance in the urine of women.27, 28,29 This is also

244

consistent with levels observed in hospital wastewater in our study, and those previously

245

reported in WTPs in Italy and China.6, 15

246

Although detected with lower frequency, concentrations of ADR (N.D.-1900 ng/L) and

247

TADR (N.D.-820 ng/L) were higher than those of ADD (N.D.-360 ng/g) and TTR (N.D.-390

248

ng/L). The concentration of TTR was comparable to that found in the WTP in Beijing, China,15

249

while ADD, ADR and TADR were lower

15

but similar to South East Queensland, Australia.30


12

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250

The highest concentration observed among the 14 steroids detected was ADR (1900 ng/L), which

251

is similar to that reported in the WTP influents collected in Beijing, China, where ADR was also

252

found to be predominant.15

253

Among the 15 progestins and metabolites in the WTP influent (Fig. 3, Table S4), P4,

254

17α-HPA and 21α-HPA, were detected at concentrations of N.D.-21 ng/L, N.D.-23 ng/L and

255

N.D.-56 ng/L, respectively. Of the two synthetic progestins, MPA ranged from N.D.-120 ng/L

256

and MTA was detected only once in influent (11 ng/L). Concentrations of P4 were lower than

257

those in WTPs in Beijing, China, 15 and comparable to another Swiss WTP.10 Concentrations of

258

17α-HPA and 21α-HPA were slightly higher than those reported in the WTP from Beijing, China,

259

15

260

were generally lower than in Beijing, China, except for one influent sample of WTP Niederdorf,

261

where we found 120 ng/L of MPA. The P4 metabolite PD and 17OH-∆5P were detected in the

262

influents with concentrations of N.D.- 1200 ng/L and N.D.- 260 ng/L, respectively, which is in

263

the same range as detected previously in influents of Swiss WTPs (N.D.- 400 ng/L and 51-2600

264

ng/L).19 In the corresponding 24 h composite effluent samples of the same WTPs, only 6

265

steroids were detected, including three estrogens (E1, E2 and E3), one androgen (ADD), and two

266

progestins (P4, MPA) (Fig. 3, Table S5). Highest levels of ADD, E1 and P4 were 1.6 ng/L, 8.7

267

ng/L and 0.23 ng/L in 2016, and P4 reached 3.3 ng/L in 2015. These three steroids were most

268

frequently detected with frequencies of 92%, 67% and 33%, respectively. They were also

269

dominant but occurred at lower levels in WTPs in Beijing, China.15

but comparable to that in another Swiss WTP.10 The concentrations of synthetic progestins

270

In the WTPs of the canton Zürich area (WTP Rüti, Wetzikon, Hinwil, Bubikon), P4 was

271

always detected in influents and effluents, with a maximal concentration of 28 ng/L and 3.7 ng/L,

272

respectively, while TTR occurred up to 82 ng/L in the influent and 1.3 ng/L in the effluent. COR


13


273

Page 14 of 33

was detected only once in influent (57 ng/L; Table S6).

274 275

Occurrences of Steroid Hormones in Rivers receiving WTP effluents. We also alalzed river

276

water samples receiving WTP effluent in the regions of the cantons Basel (Tables S8, S9) and

277

Zürich (Table S10). As for 15 water samples collected from 4 rivers on February, March and

278

April 2016, out of 33 steroids analyzed, only ADD, E1, E2 and P4 were detected with maximum

279

concentrations of 1.5, 3.7, 1.2 and 0.21 ng/L, respectively, and at frequencies of 87%, 100%,

280

13% and 40%, respectively (Figure 4, Table S8). This is much lower than levels found in the

281

urban river in Beijing, China (E1: 0.38-8.0 ng/L, ADD: 99 ng/L and P4: 199 ng/L).11 Similar to

282

the WTP effluents in the area of the river Birs, ADD, E1 and P4 were predominant in river water.

283

In samples collected in 2015 at the same river sites and river Rhine, COR, 21α-HPA, MTA,

284

NTD and P4 were detected out of 20 steroids analyzed at concentrations of up to 32 ng/L, 28

285

ng/L, 4.6 ng/L, 9.7 ng/L and 0.34 ng/L, respectively (Table S9). P4 concentrations were

286

comparable to 2016. In the 24 h composite samples from the Rhine river, COR, 21α-HPA, MTA

287

and P4 were detected. P4 was always present between 0.11 to 1.1 ng/L, while COR and 21α-HPA

288

were up to 1.1 ng/L and 19 ng/L, respectively (Table S9). In the three rivers (Aabach, Schwarz,

289

Wildbach) of the canton Zürich area that receive inputs from municipal WTPs only TTR and P4

290

were detected of up to 0.81 and 5.4 ng/L (Table S10). P4 was always detected between 0.44 and

291

5.4 ng/L. The comparison between WTP effluents and receiving river waters downstream

292

indicated that the WTP effluents were major contributors to rivers flows (Tables S7, S8, S9,

293

S10).

294 295

Removal Efficiencies of WTP with Activated Sludge Treatment. The removal was


14

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296

estimated by comparing concentrations of individual steroids in flow-proportional 24 h influent

297

and effluent samples. High removal efficiency of androgens, estrogens and progestogens were

298

observed in the WTPs with mean removal of 91±9.6% for ADD, 99±1.3% for ADR, 99±0.7% for

299

TADR, 99±1.8% for TTR, 95±8.3% for E1, 96±6.9% for E2, 100±0.2% for E3, 100±0.1% for

300

17α-HPA, 94±12% for P4, 100±0.03% for 21α-HPA, 93±8.7% for MPA, 99.6% for MTA,

301

100±0.2% for 17OH-∆5P and 100±0.04% for PD. Similar high removal efficiencies have been

302

reported for androgens and progestins in Chinese WTPs with activated sludge treatment.15 The

303

removal efficiencies of estrogens were higher than those reported in previous studies that ranged

304

between 22% and 100%.31

305

In the half-years composite influent and effluent samples of the municipal WTP of the

306

city of Basel (Prorheno AG) only TTR and P4 were detected. The concentration of TTR in the

307

influent was 68 ng/L (collected from January to June), and 24 ng/L (collected from July to

308

December) (Table S7). The P4 concentration in the influent was 13 ng/L (mixture from January

309

to June), and 8.4 ng/L (from July to December). In the corresponding WTP effluent samples, the

310

concentration of TTR was 0.8 ng/L (January to June) and 0.7 ng/L (July to December), and P4

311

concentrations were 1.5 ng/L (January to June) and 1.2 ng/L (July to December). The removal

312

efficiencies were 97% for TTR and 86% for P4, which was similar to those found in the other

313

municipal WTPs. However, storage may have resulted in a possible loss of steroids.6, 32

314 315

Removal efficiencies in WTPs with ozonation and pulverized activated carbon

316

treatment. Two WTPs were studied with advanced treatment, including ozonation in WTP

317

Neugut, Dübendorf, and activated carbon treatment in WTP Herisau. Ozonation (Table S11) and

318

pulverized activated carbon treatment (Table S12), were compared to assess further removal


15


Page 16 of 33

319

efficiencies by advanced treatment (Figure 4). In the influents of WTP Dübendorf and WTP

320

Herisau, 12 and 8 steroids, respectively, were detected. Among them ADR and TADR showed

321

the highest concentrations of 2000 ng/L and 1200 ng/L, respectively. The steroids concentrations

322

were similar as in the other municipal WTPs (Tables S4, S6). Moreover, mifepristone (MFST)

323

occurred once with a maximum concentrations of 10 ng/L in WTP Dübendorf. In the effluent of

324

WTP Dübendorf, only ADD, E1 and P4 occurred with concentrations up to 0.9, 0.3 and 1.0 ng/L,

325

respectively, and in WTP Herisau up to 0.4, 0.3 and 0.87 ng/L, respectively.

326

The comparison of influent and effluent concentrations shows that ozonation was

327

effective in further removing remaining concentrations of ADD, ADR, TADR, TTR, E1, 17α-

328

HPA and P4 after activated sludge treatment to concentrations lower than 1 ng/L (Figure 4, Table

329

S11). Removal efficiencies were up to 99% when their residues were higher than 1 ng/L. At

330

lower concentrations, both treatments did not further improve removal.

331 332

Responses of Zebrafish Embryos. To investigate potential ecotoxicological implications

333

of the steroids found in our study, we measured multi end-point responses of zebrafish eleuthero-

334

embryos, including development, behavior and transcriptional responses of target genes, in

335

response to two widely detected steroids, PD and 17OH-∆5P, and two artificial mixtures

336

representing the here found concentrations in wastewater and river water, respectively (Figures

337

4, 5). The measured concentrations in exposure water were lower than nominal with mean

338

concentrations (between 0 h and 24 h) of 429 ng/L PD, and 162 ng/L 17OH-∆5P, respectively,

339

due to sorption to surfaces and uptake by embryos. This decrease is similar as in our previous

340

studies. 12,13 The river water mixture (mix-R) consisted of P4 and ADD with mean concentrations

341

of 0.88 and 1.48 ng/L, respectively. The wastewater effluent mixture (Mix-W) consisted of P4,


16

Page 17 of 33


342

TTR, PRE, ADD, ADR, E2, E1, 17OH-∆5P and PD in the range between 5.35 and 341 ng/L each

343

(Table S13). The frequency of spontaneous muscle contractions of 24 hpf embryos showed a

344

significant decrease (47%) in response to Mix-W (Figure 5A). Similarly, heartbeat rates of 36

345

hpf embryos were induced by 19% (Figure 5B). Spontaneous contractions and heartbeat rates

346

displayed similar changes in response to Mix-R, although not significantly, with a 23% decrease

347

and 6.0% increase, respectively (Figure 5A, 5B). These results were consistent with previous

348

reports that glucocorticoids, like prednisolone and fludrocortisone acetate, can reduce zebrafish

349

embryo spontaneous contraction and increase heartbeat rate.

350

were also reported to affect heart structure and function of zebrafish embryos.34, 35 Injection of

351

cortisol into embryos caused heart deformities, such as pericardial edema and malformed

352

chambers, and suppressed essential cardiac gene expression.34 Though no significant

353

physiological defects of the hearts were observed in the present study, the results suggest

354

disruptions of embryonic cardiogenesis in the Mix-W exposure.

20, 33

Dexamethasone and cortisol

355

Transcriptional responses of four crucial genes (pepck1, g6pca, socs3 and slco2a1) either

356

involved in gluconeogenesis or immune responses were further measured, as they showed

357

changes upon corticosteroids exposure.20, 35, 36 Of these genes, socs3 was down-regulated, and

358

significant responses were observed upon both Mix-W and Mix-R exposures (Figure 5D). Socs3

359

encodes a member of the STAT-induced STAT inhibitor, which is a negative regulator of

360

cytokine signaling. It can be induced by cytokines and participates in the inhibition of JAK2

361

kinase activity.37 Decreases of socs3 further suggests that corticosteroid effects existed in these

362

mixtures. On the other hand, hatching success at 48 hpf and swimming behavior (activity) at 72

363

hpf were not significantly affected (Figure 5C, S1).

364

Besides, transcriptional responses of target genes involved in steroidogenesis and related


17


Page 18 of 33

365

nuclear receptors were measured (Figure 6A). Of these, six genes, including erα, er2β, ar, pr,

366

hsd11b2 and hsd17b3, were significantly down regulated in response to Mix-W exposure (Figure

367

S2). Among them, erα and pr showed the most pronounced changes, with fold-changes of up to

368

1.6 times. Interestingly, two steroidogenesis genes, hsd11b2 and hsd17b3, were significantly

369

down regulated in response to Mix-R exposure (Figure 6B), though no obvious physiological

370

effects were observed. Any of the genes were significantly up-regulated.

371

These results are consistent with previous reports that steroids, such as estrogens and

372

progestins, can alter expression of genes in hypothalamic–pituitary–gonadal axis (HPG axis) at

373

environmental relevant levels (ng/L).37, 38 The artificial mixture of river water contained 0.88

374

ng/L P4 and 1.48 ng/L ADD, respectively. The significant transcriptional changes in

375

steroidogenesis enzymes under Mix-R exposure suggests a joint action of the steroids in the

376

aquatic environment, which should further be substantiated in forthoming studies.

377

Responses on the embryonic development upon exposure to 429 ng/L PD and 162 ng/L

378

17OH-∆5P were also determined but most end-points, including spontaneous contractions,

379

heartbeat rate, behaviors and target genes involved in gluconeogenesis and immune responses

380

were not significantly changed (Figure S4). Of the HPG axis genes, expression of erα and er2β

381

displayed significant but slight changes (Figure 6). This suggest that PD and 17OH-∆5P at

382

environmental concentrations may have negligible effects on early fish development.

383

In conclusion, our analysis showed the presence of considerable levels of several

384

estrogens, androgens, progestins and metabolites in hospital wastewater and raw wastewater, but

385

low concentrations in treated wastewater and receiving river waters in Switzerland. P4 was the

386

most detected progestin with concentrations usually below 1 ng/L. Steroids were efficiently

387

removed in WTPs (93-100%) with conventional activated sludge treatment, and advanced


18

Page 19 of 33


388

treatment technologies further decreased remaining steroids. Furthermore, zebrafish embryo

389

exposures demonstrated negligible effects of environmental levels of widely detected PD and

390

17OH-∆5P, while mixtures that mimic wastewater and river water composition affected

391

zebrafish embryo development and led to alteration of steroidogenesis gene transcripts at ng/L.

392

Further studies are needed to confirm the low occurrence of steroids in the environment and to

393

assess the ecotoxicological consequences of steroid mixtures to aquatic organisms.

394 395

Supporting information

396

Details of chemical analysis; abbreviations of target compounds; list of samples collected (Table

397

S1); method detection limits and recoveries (Table S2); concentrations of steroid hormones in

398

hospital wastewater (Table S3), WTPs influent and effluent wastewater (Table S4-S7), river

399

samples (Table S8-S10), WTPs with advanced treatment technologies (Table S11-S12), and

400

samples collected during Zebrafish embryo exposure (Table S13). This material is available free

401

of charge via the internet at http://pubs.acs.org.

402 403 404 405

Acknowledgments

406

(canton Basel-Stadt), Antje Langbein (FHNW), Ulrich Muggli and Mark Morley (WTP

407

Wetzikon), Markus Vock (WTP Birsfelden), Kathrin Halbeisen and Ursula Garbani and all other

408

staff in WTP Herisau, Nathalie Hubaux and all other staff in WTP Neugut (Dübendorf), for help

409

in samling and providing WTP samples, Fabian Honegger and Daniel Bollhalder, for help in

410

sampling at GZO Spital Wetzikon, Francsco Di Benedetto and Manuel Grieder for sampling at

411

hospital Basel-Land and Liestal. The study was funded by the Swiss National Science

412

Foundation (contract no. 310030_141040 and 310030_169205 to K.F.) and the Swiss Federal Office

413

for the Environment (contract no. 00.0296.PZ/O483-0261 to K.F.).

We thank Adrian Auckenthaler, Martin Huser (canton Basel-Land), Reto Dolf and Jan Mazacek

414


19


415 416

REFERENCES

417

(1) Mills, L. J.; Chichester, C. Review of evidence: are endocrine-disrupting chemicals in the

418

aquatic environment impacting fish populations? Sci. Total Environ. 2005, 343 (1), 1-34.

419

(2) Fent, K. Progestins as endocrine disrupters in aquatic ecosystems: concentrations, effects and

420

risk assessment. Environ. Int. 2015, 84, 115-130.

421

(3) Sumpter, J. P.; Johnson, A. C. Lessons from endocrine disruption and their application to

422

other issues concerning trace organics in the aquatic environment. Environ. Sci. Technol. 2005,

423

39 (12), 4321-4332.

424

(4) Jobling, S.; Williams, R.; Johnson, A.; Taylor, A.; Gross-Sorokin, M.; Nolan, M.; Tyler, C. R.;

425

van Aerle, R.; Santos, E.; Brighty, G. Predicted exposures to steroid estrogens in U.K. rivers

426

correlate with widespread sexual disruption in wild fish populations. Environ. Health Perspect.

427

2006, 114 Suppl 1, 32-39.

428

(5) Tyler, C. R.; Jobling, S. Roach, sex, and gender-bending chemicals: The feminization of wild

429

fish in English rivers. Bioscience 2008, 58 (11), 1051-1059.

430

(6) Baronti, C.; Curini, R.; D'Ascenzo, G.; Di Corcia, A.; Gentili, A.; Samperi, R. Monitoring

431

natural and synthetic estrogens at activated sludge sewage treatment plants and in a receiving

432

river water. Environ. Sci. Technol. 2000, 34 (24), 5059-5066.

433

(7) Belfroid, A.; Van der Horst, A.; Vethaak, A.; Schäfer, A.; Rijs, G.; Wegener, J.; Cofino, W.

434

Analysis and occurrence of estrogenic hormones and their glucuronides in surface water and

435

waste water in The Netherlands. Sci. Total Environ. 1999, 225 (1), 101-108.

436

(8) Johnson, A.; Aerni, H.-R.; Gerritsen, A.; Gibert, M.; Giger, W.; Hylland, K.; Jürgens, M.;

437

Nakari, T.; Pickering, A.; Suter, M.-F. Comparing steroid estrogen, and nonylphenol content

438

across a range of European sewage plants with different treatment and management practices.

439

Water Res. 2005, 39 (1), 47-58.

440

(9) Vulliet, E.; Cren-Olivé, C. Screening of pharmaceuticals and hormones at the regional scale,

441

in surface and groundwaters intended to human consumption. Environ. Pollut. 2011, 159 (10),

442

2929-2934.

443

(10) Ammann, A. A.; Macikova, P.; Groh, K. J.; Schirmer, K.; Suter, M. J. LC-MS/MS

444

determination of potential endocrine disruptors of cortico signalling in rivers and wastewaters.

445

Anal. Bioanal.Chem. 2014, 406 (29), 7653-7665.

446

(11) Chang, H.; Wan, Y.; Hu, J. Determination and source apportionment of five classes of


20

Page 20 of 33

Page 21 of 33


447

steroid hormones in urban rivers. Environ. Sci. Technol. 2009, 43 (20), 7691-7698.

448

(12) Liu, Z.-h.; Kanjo, Y.; Mizutani, S. Urinary excretion rates of natural estrogens and

449

androgens from humans, and their occurrence and fate in the environment: a review. Sci. Total

450

Environ. 2009, 407 (18), 4975-4985.

451

(13) Labadie, P.; Budzinski, H. Determination of steroidal hormone profiles along the Jalle

452

d'Eysines River (near Bordeaux, France). Environ. Sci. Technol. 2005, 39 (14), 5113-5120.

453

(14) Liu, S.; Chen, H.; Xu, X.-R.; Liu, S.-S.; Sun, K.-F.; Zhao, J.-L.; Ying, G.-G. Steroids in

454

marine aquaculture farms surrounding Hailing Island, South China: occurrence,

455

bioconcentration, and human dietary exposure. Sci. Total Environ. 2015, 502, 400-407.

456

(15) Chang, H.; Wan, Y.; Wu, S.; Fan, Z.; Hu, J. Occurrence of androgens and progestogens in

457

wastewater treatment plants and receiving river waters: Comparison to estrogens. Water Res.

458

2011, 45 (2), 732-740.

459

(16) Pauwels, B.; Noppe, H.; De Brabander, H.; Verstraete, W. Comparison of steroid hormone

460

concentrations in domestic and hospital wastewater treatment plants. J. Environ. Eng. 2008, 134

461

(11), 933-936.

462

(17) Wudy, S.; Hartmann, M. Gas chromatography-mass spectrometry profiling of steroids in

463

times of molecular biology. Hormone and metabolic research 2004, 36 (06), 415-422.

464

(18) Besse, J.-P.; Garric, J. Progestagens for human use, exposure and hazard assessment for the

465

aquatic environment. Environ. Pollut. 2009, 157 (12), 3485-3494.

466

(19) Zhang, K.; Fent, K. Determination of two progestin metabolites (17α-hydroxypregnanolone

467

and pregnanediol) and different classes of steroids (androgens, estrogens, corticosteroids and

468

progestins) in rivers and wastewaters by high-performance liquid chromatography-tandem mass

469

spectrometry (HPLC-MS/MS). Environ. Sci. Technol., submitted.

470

(20) Zhao, Y.; Zhang, K.; Fent, K. Corticosteroid Fludrocortisone Acetate Targets Multiple End

471

Points in Zebrafish (Danio rerio) at Low Concentrations. Environ. Sci. Technol. 2016, 50 (18),

472

10245-10254.

473

(21) Siegenthaler, P. F.; Bain, P.; Riva, F.; Fent, K. Effects of antiandrogenic progestins,

474

chlormadinone and cyproterone acetate, and the estrogen 17α-ethinylestradiol (EE2), and their

475

mixtures: Transactivation with human and rainbowfish hormone receptors and transcriptional

476

effects in zebrafish (Danio rerio) eleuthero-embryos. Aquat. Toxicol. 2017, 182, 142-162.

477

(22) Livak, K. J.; Schmittgen, T. D. Analysis of relative gene expression data using real-time


21


478

quantitative PCR and the 2− ∆∆CT method. methods 2001, 25 (4), 402-408.

479

(23) Hauser, B.; Deschner, T.; Boesch, C. Development of a liquid chromatography–tandem

480

mass spectrometry method for the determination of 23 endogenous steroids in small quantities of

481

primate urine. J. Chromatogr. B 2008, 862 (1), 100-112.

482

(24) Preedy, J. R.; Aitken, E. H. The determination of estrone, estradiol-17β, and estriol in urine

483

and plasma with column partition chromatography. J. Biol. Chem. 1961, 236 (5), 1300-1311.

484

(25) Van Renterghem, P.; Van Eenoo, P.; Geyer, H.; Schänzer, W.; Delbeke, F. T. Reference

485

ranges for urinary concentrations and ratios of endogenous steroids, which can be used as

486

markers for steroid misuse, in a Caucasian population of athletes. Steroids 2010, 75 (2), 154-163.

487

(26) Nagarnaik, P. M. Source Characterization and Pretreatment Evaluation of Pharmaceuticals

488

and Personal Care Products in Healthcare Facility Wastewater. Texas A&M University, 2011.

489

(27) Fotsis, T.; Adlercreutz, H. The multicomponent analysis of estrogens in urine by ion

490

exchange chromatography and GC-MS—I. Quantitation of estrogens after initial hydrolysis of

491

conjugates. Journal of Steroid Biochemistry 1987, 28 (2), 203-213.

492

(28) Fotsis, T.; Järvenpää, P.; Adlercreutz, H. Purification of urine for quantification of the

493

complete estrogen profile. Journal of Steroid Biochemistry 1980, 12, 503-508.

494

(29) Andreolini, F.; Borra, C.; Caccamo, F.; Di Corcia, A.; Samperi, R. Estrogen conjugates in

495

late-pregnancy fluids: extraction and group separation by a graphitized carbon black cartridge

496

and quantification by high-performance liquid chromatography. Anal. Chem. 1987, 59 (13),

497

1720-1725.

498

(30) Tan, B. L.; Hawker, D. W.; Müller, J. F.; Leusch, F. D.; Tremblay, L. A.; Chapman, H. F.

499

Comprehensive study of endocrine disrupting compounds using grab and passive sampling at

500

selected wastewater treatment plants in South East Queensland, Australia. Environ. Int. 2007, 33

501

(5), 654-669.

502

(31) Liu, Z.-h.; Kanjo, Y.; Mizutani, S. Removal mechanisms for endocrine disrupting

503

compounds (EDCs) in wastewater treatment—physical means, biodegradation, and chemical

504

advanced oxidation: a review. Sci. Total Environ. 2009, 407 (2), 731-748.

505

(32) Vanderford, B. J.; Mawhinney, D. B.; Trenholm, R. A.; Zeigler-Holady, J. C.; Snyder, S. A.

506

Assessment of sample preservation techniques for pharmaceuticals, personal care products, and

507

steroids in surface and drinking water. Anal. Bioanal.Chem. 2011, 399 (6), 2227-2234.

508

(33) McNeil, P. L.; Nebot, C.; Sloman, K. A. Physiological and behavioral effects of exposure to


22

Page 22 of 33

Page 23 of 33


509

environmentally relevant concentrations of prednisolone during zebrafish (Danio rerio)

510

embryogenesis. Environ. Sci. Technol. 2016, 50 (10), 5294-5304.

511

(34) Nesan, D.; Vijayan, M. M. Embryo exposure to elevated cortisol level leads to cardiac

512

performance dysfunction in zebrafish. Molecular and cellular endocrinology 2012, 363 (1), 85-

513

91.

514

(35) Wilson, K.; Baily, J.; Tucker, C.; Matrone, G.; Vass, S.; Moran, C.; Chapman, K.; Mullins,

515

J.; Kenyon, C.; Hadoke, P. Early-life perturbations in glucocorticoid activity impacts on the

516

structure, function and molecular composition of the adult zebrafish (Danio rerio) heart.

517

Molecular and cellular endocrinology 2015, 414, 120-131.

518

(36) Kugathas, S.; Runnalls, T. J.; Sumpter, J. P. Metabolic and reproductive effects of relatively

519

low concentrations of beclomethasone dipropionate, a synthetic glucocorticoid, on fathead

520

minnows. Environ. Sci. Technol. 2013, 47 (16), 9487-9495.

521

(37) Chen, Q.; Jia, A.; Snyder, S. A.; Gong, Z.; Lam, S. H. Glucocorticoid activity detected by in

522

vivo zebrafish assay and in vitro glucocorticoid receptor bioassay at environmental relevant

523

concentrations. Chemosphere 2016, 144, 1162-1169.

524

(38) Yin, Y.; Liu, W.; Dai, Y. SOCS3 and its role in associated diseases. Human immunology

525

2015, 76 (10), 775-780.

526 527


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528

Figure legends

529

Figure 1. Map of sampling area of Basel and canton Zürich, Switzerland. (A) Sampling sites and

530

WTPs in Basel. (P: Prorheno AG, B0: Therwil (Birsig), B2: Birsfelden (Birs II), E1: Sissach

531

(Ergolz I), E2: Füllinsdorf (Ergolz II), F2: Niederdorf (Frenke II), F3: Bubendorf (Frenke III),

532

R0: Rhine river R1: River Birs, R2: River Birs, R3: River Lüssel, R4: River Lucelle_, R5: River

533

Ergolz_). (B) Sampling sites and WTPs in canton Zürich (W: WTP Wetzikon, H: WTP Hinwil,

534

B: WTP Bubikon, R7:River Aabach, R8: River Wildbach, R9:River Schwarz).

535 536

Figure 2. Concentrations of different steroids higher than 5 ng/L in the waste water samples

537

collected from GZO Spital Wetzikon (Wetzikon city hospital) at different time points. PD:

538

pregnanediol, 17OH-Δ5P: hydroxypregnanolone, P4:progesterone, EE2:ethinyl estradiol,

539

E3:estriol, E2:17β-estradiol, E1:estrone, TTR:testosterone, TADR: trans-androsterone, ADR:

540

androsterone, ADD: androstenedione.

541

Figure 3. Concentration of steroids in influents and effluents of six WTPs in Basel-Land,

542

Switzerland, ((B0: Therwil (Birsig), B2: Birsfelden (Birs II), E1: Sissach (Ergolz I), E2:

543

Fuellinsdorf (Ergolz II), F2: Niederdorf (Frenke II), F3: Bubendorf (Frenke III) ) and river

544

samples collected from River Birs, Lüssel, Lucelle and Ergolz (R1-R5) receiving inputs from the

545

WTPs in Basel, Switzerland. Abbreviations see Materials section. Day 1: March 10, 2016, Day

546

2: June 28, 2016, Day 3: February 3, 2016, Day 4: March 3, 2016, Day 5: April 5, 2016. PD:

547

pregnanediol, 17OH-Δ5P: hydroxypregnanolone, MTA: megestrol acetate, MPA:

548

medroxyprogesterone acetate, P4:progesterone, 21α-HPA: 21α-hydroxyprogesterone, 17α-


24

Page 24 of 33

Page 25 of 33


549

HPA: 17α-hydroxyprogesterone, E3:estriol, E2:17β-estradiol, E1:estrone, TTR:testosterone,

550

TADR: trans-androsterone, ADR: androsterone, ADD: androstenedione.

551

Figure 4. Concentration of steroids in wastewater samples collected at different treatment steps

552

of WTP Neugut, Dübendorf, consisting of an ozonation step, and WTP Herisau with pulverized

553

activated carbon treatment. Abbreviations see Materials section. Day 1: March 24, 2016, Day 1:

554

March 28, 2016, Day 3: July 5, 2016. PD: pregnanediol, 17OH-Δ5P: hydroxypregnanolone,

555

MTA: megestrol acetate, MPA: medroxyprogesterone acetate, P4: progesterone, 21α-HPA: 21

556

α-hydroxyprogesterone, 17α-HPA: 17α-hydroxyprogesterone, E3:estriol, E2:17β-estradiol,

557

E1:estrone, TTR:testosterone, TADR: trans-androsterone, ADR: androsterone, ADD:

558

androstenedione.

559

Figure 5. Eleuthero-embryo development, swimming behavior and gene transcript levels upon

560

embryonic exposure to 429 ng/L PD, 162 ng/L 17OH-∆5P and artificial mixtures of wastewater

561

(Mix-W) and river water (Mix-R). Mix-R consisted of P4 and ADD with mean concentrations of

562

0.88 and 1.48 ng/L, respectively. Mix-W consisted of P4, TTR, PRE, ADD, ADR, E2, E1,

563

17OH-∆5P and PD in the range between 5.35 and 341 ng/L for each compound. (A) Contraction

564

rate of embryos at 24 hpf (n = 24). (B) Heartbeat rate of embryos at 36 hpf (n = 24). (C)

565

Swimming behavior (activity) of eleuthero-embryos at 72 hpf (n = 8). (D) Transcriptional

566

alteration of socs3 (n = 4 replicates) after 48h of exposure. Asterisks* indicates p-value

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