Accepted Manuscript Title: Occurrence and removal of sulfonamide antibiotics and antibiotic resistance genes in conventional and advanced drinking water treatment processes Authors: Yaru Hu, Lei Jiang, Tianyang Zhang, Lei Jin, Qi Han, Dong Zhang, Kuangfei Lin, Changzheng Cui PII: DOI: Reference:
S0304-3894(18)30690-3 https://doi.org/10.1016/j.jhazmat.2018.08.012 HAZMAT 19636
To appear in:
Journal of Hazardous Materials
Received date: Revised date: Accepted date:
14-3-2018 25-7-2018 5-8-2018
Please cite this article as: Hu Y, Jiang L, Zhang T, Jin L, Han Q, Zhang D, Lin K, Cui C, Occurrence and removal of sulfonamide antibiotics and antibiotic resistance genes in conventional and advanced drinking water treatment processes, Journal of Hazardous Materials (2018), https://doi.org/10.1016/j.jhazmat.2018.08.012 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. 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.
Occurrence and removal of sulfonamide antibiotics and antibiotic resistance genes in conventional and advanced drinking water treatment processes
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Yaru Hu1, Lei Jiang2, Tianyang Zhang1,3, Lei Jin2, Qi Han1, Dong Zhang2,
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Kuangfei Lin1, Changzheng Cui*1,3
1. State Environmental Protection Key Laboratory of Environmental Risk Assessment and Control on Chemical Process, School of Resources and Environmental
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Engineering, East China University of Science and Technology, Shanghai, China,
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200237
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2. National Engineering Research Center of Urban Water Resources, Shanghai, China,
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200082
3. Shanghai Institute of Pollution Control and Ecological Security, Shanghai, 200092,
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PR China
*Corresponding author, Changzheng Cui, Tel: +86 21 64253988; Fax: +86 21
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64253988; E-mail:
[email protected]
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Graphical abstract
Removal of SAs and genes
Raw water 100
2.5
80
2.0
CL
PCL→SF
60
40
1.0
April
July
Advanced process
Log reduction of genes
CL
60
40
sul1
1.0
0
April
July
September
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TMP SAM SG SDZ SMR SMZ STZ SMTSMX SFX SCP SMD SAT
SAs, sul ARGs and intI1 were detected in two different DWTPs.
SMX was the most abundant SAs in both raw water and finished water in two
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DWTPs.
Advanced process gained an advantage over conventional process in
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eliminating SAs.
The residual sul ARGs and intI1 in finished water remained as 102-104 gene abundance/mL. Conventional treatment units was more stable in removing sul ARGs and intI1.
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2
intI1
1.5
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Highlights
sul2
0.5
20
0
November
2.0
80
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GAC
Removal efficiency of SAs (%)
OD
September
2.5
16S rDNA rRNA
FLO+SF
intI1
1.5
0
TMP SAM SG SDZ SMR SMZ STZ SMTSMX SFX SCP SMD SAT
100
PO
sul2
0.5
20
0
sul1
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Conventional process
Log reduction of genes
13 SAs sul1, sul2 intI1
Removal efficiency of SAs (%)
16S rDNA rRNA
November
Abstract Sulfonamides (SAs) and sul antibiotic resistance genes (ARGs) have been extensively detected in drinking water sources and warrant further studies on the removal of them in different drinking water treatment processes (DWTPS). The
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prevalence of 13 SAs, sul1, sul2 and class I integrase gene intI1 in conventional and
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advanced processes was investigated using HPLC-MS/MS and real-time quantitative
PCR (qPCR), respectively. The most abundant SA was sulfamethoxazole, with the maximum concentration of 67.27 ng/L. High concentration of sulfamethoxazole was
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also measured in finished water in both conventional (22.05 ng/L) and advanced
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(11.24 ng/L) processes. Overall, the removal efficiency of advanced process for each
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SA was higher than that of conventional process, except for sulfameter. The absolute concentrations of sul1, sul2 and intI1 in raw water ranged from 1.8×103 to 2.4×105
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gene abundance/mL. After treatment, the residual sul ARGs and intI1 in finished
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water still remained at 102 - 104 gene abundance/mL. Conventional treatment units,
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including flocculation/sedimentation/sand filtration, played a more important role in removing sul1, sul2 and intI1 than oxidation (chlorination or ozonation) and granular activated carbon filtration treatments. Based on this work, more investigations are
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needed to help improve the removal of both antibiotics and ARGs in DWTPS.
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Keywords: sulfonamides (SAs); antibiotic resistance genes (ARGs); drinking water treatment plants (DWTPs); conventional and advanced processes, removal efficiency
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1 Introduction
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SAs, a class of antimicrobials with amino benzenesulfonyl structure and target
dihydropteroate synthetase [1], have larger production and consumption in China than in other countries [2]. With the low octanol-water partition coefficients (logKow) and
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high emission rates, SAs have high mobility and are frequently detected with high
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concentrations in the water environment [3,4]. For example, SAs were detected in 83 -
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94% of the samples in Haihe River in China, with the highest concentrations of 210 385 ng/L [5]. Several studies had also confirmed the existence of SAs in drinking
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water sources in China (Anhui: 0.2 - 12.5 ng/L [6]; Nanjing: 4.4 - 78.8 ng/L [7]), even
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1,840 ng/L sulfamethoxazole (SMX) was detected in groundwater in Taiwan [8].
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What’s more, in the United States, sulfamethazine (SMZ) and SMX were detected in groundwater in 18 states with concentrations ranging from 360 to 1,110 ng/L [9]. In Vietnam, the concentration of SMX in the city canal even reached 4,330 ng/L [10].
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The residual SAs in aquatic environments not only pose a threat to aquatic creatures, but also produce a selective pressure for environmental microorganisms and contribute to the problems of antibiotic resistance in microorganisms [11,12]. 4
As an emerging contaminant, ARGs have attracted global attention worldwide. They are usually found coexisting with the mobile genetic elements, such as integrons, plasmids, which would promote the horizontal gene transfer [13,14]. Once these ARGs transmitted into pathogenic bacteria, its expression in harmful pathogens will
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undermine our ability to treat infectious diseases and exert a serious threat to human
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health [14]. Among the common four sul ARGs (sul1, sul2, sul3, sulA), sul1 and sul2
are two main sul ARGs reported in recent studies [3,15]. For example, the sul1 and sul2 concentrations in Huangpu River were both more than 105 gene abundance/mL
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[16]. In Haihe River, sul ARGs were 100% detected in water samples with absolute
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concentrations of 107 - 108 gene abundance/mL [3]. Su et al. [17] isolated 3,456 E.
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coli from Dongjiang River (Guangzhou) and 89.1% of them were resistant to at least three antibiotics and the frequency of sul ARGs reached 89.2%. Besides, the pollution
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of sul ARGs was also detected in Beijiang River, Guangzhou [18], the Manasi River
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Basin, Xinjiang [19], and Pearl River [15]. All of these studies provided the evidence
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for the pollution of the aquatic environment by SAs and sul ARGs. Recently, researchers have started to pay attention to the removal of antibiotics
and ARGs in water treatment process [20]. Gaffney et al. [21] found that 65% of SAs
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could be removed during the chlorination process. The bio-treatment process achieved 77% removal of antibiotics and was efficient in removing ARGs in veterinary hospital wastewater [22]. Other methods, such as ultraviolet (UV) irradiation [23], 5
UV/chlorination [24,25], and ozonation processes [26] were also proved to be effective in eliminating SAs and ARGs (0.80 - 2.28 log for ARGs). However, these studies only focused on the removal efficiencies of individual water treatment units
in the whole DWTPS should also be paid high attention to.
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and the target samples were all wastewater. Therefore, the removal of SAs and ARGs
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In this study, the occurrence and removal of 13 SAs, two sul ARGs (sul1, sul2) and class I integrase gene intI1 were investigated in water samples from the different units of two DWTPs with conventional and advanced treatment processes. The 13
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SAs were detected by SPE-HPLC/MS/MS and ARGs were quantified by qualitative
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polymerase chain reaction (qPCR), respectively. The results will provide a basis to
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understand the prevalence of SAs and ARGs in the DWTPS. The comparison between the conventional and advanced processes will benefit the DWTPs for optimizing the
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treatment methods to eliminate these contaminants simultaneously.
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2 Materials and methods
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2.1 Standards and reagents Standards of 13 SAs including trimethoprim (TMP), sulfanilamide (SAM),
sulfaguanidine (SG), sulfadiazine (SDZ), sulfamerazine (SMR), SMZ, sulfathiazole
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(STZ), sulfamethizole (SMT), SMX, sulfisoxazole (SFX), sulfachloropyridazine (SCP), sulfameter (SMD), sulfadimethoxine (SAT), and SMX isotope marker (SMX-13C6) were purchased from Sigma (USA). Hydrochloric acid and edetate 6
disodium (EDTA) were obtained from Sinopharm Chemical Reagent Co., Ltd. (China). Formic acid, methanol, and acetonitrile were HPLC-grade and purchased from Thermo Fisher Company (Fisher Scientific, USA). All standard solutions were stored at -20 °C. Ultra-pure water was prepared using a Milli-Q water purification
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system (Millipore, Bedford, MA, USA). The physicochemical properties of target
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SAs are presented in supplementary materials (Table S1). 2.2 Sample collection
All samples in this study were collected from two different DWTPs that adopted
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conventional (DWTP1) and advanced processes (DWTP2), respectively (see the
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treatment scheme in Fig. 1). These two DWTPs receive raw water from the same
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drinking water source (Yangtze River in Eastern China) but at different delivery distances (DWTP1: 21.3 km, DWTP2: 48.6 km). In DWTP1, raw water was treated
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successively by pre-chlorination → pre-/post-flocculation → sand filtration
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(PCL→SF), and chlorination (CL) before entering the distribution system (Fig. 1). In
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DWTP2, after travelling a long delivery distance (48.6 km) from the drinking water source, raw water was treated by an advanced process that involved the steps of pre-ozonation (PO), flocculation + sand filtration (FLO+SF), post-ozonation (OD),
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granular activated carbon filtration (GAC), and final chlorination. The sampling sites of two DWTPs and detailed parameters of each unit are shown in Fig. 1 and Table S2, respectively. Water samples were collected in April, July, September, and November 7
2015 and transported to the laboratory as soon as possible on ice in the dark. 2.3 Sample pretreatment Solid phase extraction (SPE) pretreatment for the analysis of SAs is mainly based on our previous study [27]. Pretreatment of water samples and solid phase
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extraction prior to HPLC-MS/MS analysis is detailed in Supporting Information (SI) Section SI-1.
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After sample collection, a 0.22 μm cellulose acetate membrane filter (SCBB-207, Anpel, Shanghai) was used to intercept the microorganisms from about 1 L water
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sample. The filters were then cut into small pieces and DNA was extracted from these
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cut filters using the Fast DNA® SPIN Kit for Soil (MP Biomedicals, USA) according
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to the manufacturer’s instructions. The quality and concentration of the extracted DNA were determined by 1.2% agarose gel electrophoresis and a NanoDrop
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spectrophotometer (NanoDrop2000, USA), respectively.
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2.4 Quantification of SAs and ARGs
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The target SAs were quantified by a high performance liquid chromatographic tandem mass spectrometry (HPLC-MS/MS, Agilent, USA) under the multiple reaction monitoring (MRM) mode and the method parameters were optimized based
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on previous studies [28]. The HPLS-MS/MS details were shown in Section SI-2 and the detail mass spectrometric parameters, recoveries, and limits of quantitation of each SA are shown in Table S3. 8
To identify the exists of target sul ARGs and intI1 genes in the studied samples and establish the quantitative standard curves, PCR detection assays were conducted based on the previous publications [3,12,16] using a Thermo Cycler (KF960, Hangzhou, China). The PCR mixtures and condition were detailed in Section SI-3.
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The identified target genes were further quantified using a LightCycle 480 II
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instrument (Roche, Swiss). The quantitative standard curves were generated as suggested [3] and using a pGEM-T Easy vector (Promega, USA) as well as the
competent cell E. coli DH5α (Tiangen, Shanghai, China), with target genes (obtained
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from the water samples) cloned into it. The qPCR mixtures and protocol condition
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were shown in Section SI-3. All qPCRs were conducted in parallel with the DNA-free
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water as negative controls. 2.5 Water quality characterization
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Some water quality parameters of the raw water and finished water of DWTPs.
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including DOC, TN, TP, NH3-N, NO2- and UV254 were analyzed routinely according
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to the Standard Methods (GB5749-2006, China) using Shimadzu TOC-VCSH analyzer (Shimadzu, Japan), ICS-1100 ion chromatography (Thermo, USA) and UV-1800 spectrophotometer (Shimadzu, Japan). Some metal ions were measured by
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ICP-MS (PerkinElmer NexION 300 X, USA). The pH was measured in the laboratory. The water quality parameters (Section SI-4) were evaluated to assess the impact of water quality conditions on the treatment processes. 9
3. Results and discussion 3.1 Occurrence and removal of SAs in conventional and advanced DWTPs 3.1.1 Occurrence of SAs in raw water and finished water The frequencies and concentrations of 13 SAs in raw water and finished water of
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two different DWTPs are shown in Table 1. During the conventional process, the
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frequencies of five SAs (TMP, SAM, SMZ, SMX and SCP) were detected as 100% in
raw water. Among these five SAs, two of them (SAM and SMZ) still showed 100% detection frequency in finished water. SMX had the highest concentrations in both
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raw water (67.27 ng/L) and finished water (22.05 ng/L) in addition to its high
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frequency of detection. As the SA synergist, TMP was detected with the second
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highest concentration level in raw water (40.86 ng/L), which was the same level of a drinking water sources in East China [27] and lower than Huangpu River [29]. The
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high levels of SMX and TMP in raw water should be attributed to the large usage in
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livestock and aquaculture in China (500 t for SMX and 238 t for TMP in 2013 [30]),
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and their chemical stability, good environmental migration ability [27,31]. Similarly, TMP, SAM, SMZ and SMX were 100% detected in raw water of
advanced process. SMX still showed the highest concentration among all the SAs in
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raw water (55.44 ng/L) and finished water (11.24 ng/L) (Table 1). SCP was not 100% detected in raw water, but its detection frequency still reached 86%. Although these two DWTPs (DWTP1: conventional process; DWTP2: advanced process) received 10
raw water from the same water source, the concentrations of SAs in raw water at these two DWTPs were different. TMP in the raw water of DWTP2 was much lower than that of DWTP1 because of the longer distribution distance from the water source to DWTP2 (48.6 km) than that to DWTP1 (21.3 km) (Fig. 1), which was more beneficial
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to biofilm generated on the transportation pipes, thus the different amounts of SAs in
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raw water would be attributed to the biological degradation of antibiotics [20,32,33]. 3.1.2 Removal of SAs in conventional process
In order to further investigate the removal of SAs in each water treatment unit,
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several sampling points were selected at these two DWTPs (Fig. 1). Fig. 2 shows the
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fractional removal of individual SAs by each treatment unit. The total height of bars
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represents the total removal efficiency of each SA. Overall, the conventional process acquired about 79% reduction of 13 SAs. The total removal rates of four SAs (TMP,
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SG, SFX and SCP) were more than 80% in the conventional process (Fig. 2(a)) and
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two SAs (SG and SFX) were completely eliminated. However, the removal
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efficiencies of SMR, STZ and SMT were less than 40% at DWTP1 (STZ removal rate was 0%). Even the removal efficiency of SMX (SA with the highest concentration)
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was 67%.
The treatment steps from PCL→SF contributed most to the removal efficiency
(average removal rate: 60.7%). This was consistent with the previous report that the treatment process of pre-ozonation/flocculation/sedimentation acquired 0 - 78% 11
reduction of PPCPs [34]. There are a pre-chlorination (0.5 - 2.0 mg/L sodium hypochlorite (NaClO), which was not sampled) and a two-stage flocculation (pre-flocculation, polyaluminium chloride and sulfate (PASC): 24 - 28 mg/L; post-flocculation, PASC: 2 mg/L + PAM: 0.11 mg/L, in Fig. 1) during PCL→SF
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process. In general, the flocculation/sedimentation process is not suitable for
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eliminating hydrophilic compounds with low logKow value and high solubility [35], and the removal rate was lower than 30% [36], so the high removal efficiencies of SAs might attribute to the PCL [37]. However, the removal of SAs during PCL would
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be affected by the organic compounds and pH [38]. For example, the removal rates of
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TMP, SCP, SMR, SMZ, and STZ by 1.0 mg/L chlorine was about 50% at pH 7.7 and
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DOC 10.7 mg/L [38], while the removal rates of TMP, SCP, SMR, SMZ, and STZ in this study ranged from 4.4% to 74.5% at pH 8.1 and DOC 2.3 mg/L.
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CL played a very limited role in eliminating most SAs, the average removal
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efficiency only reached 28%, which was the same level as a study in U.S. (32%) [39].
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Only four SAs (TMP, SAM, SMR and SMZ) showed the obvious degradation during CL (>20%). The removal rates of SDZ, SMZ was the same level as the CL unit in Germany (65% - 68%) [21]. SMR kept stable during PCL→SF process and could
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only be degraded by CL. SAM was the parent structure of other aminobenzene sulfonic acid amide of SAs and could be regenerated during oxidation and decomposition, which should be responsible for the lower removal of SAM [21,40]. 12
The low removal efficiencies of other SAs during CL might be contributed to that the natural organic matters were degraded into small molecular fractions during PCL, and competed with SAs to consume chlorine [41]. 3.1.3 Removal of SAs in advanced process
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The fractional removal efficiencies of SAs by each treatment unit during the
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advanced process are shown in Fig. 2(b). Compared with the conventional process in
Fig. 2(a), the advanced process promoted the removal efficiencies for most SAs (average removal efficiency: 84.8%). Almost all detection frequencies of SAs were
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less than 50% (except for SAM) and the average residual concentrations of SAs in
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finished water were all lower than those in the conventional process. The removal
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rates of eight SAs (TMP, SG, SDZ, SMR, SMZ, SMX, SFX and SCP) reached above 80%. Among them, SMR gained high removal efficiency, which was completely
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attributed to PO, suggesting that SMR could only be degraded by the oxidation
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processes. SCP was no longer detected after this process, which was due to the
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photodegradation [21]. Among all SAs, only two SAs (STZ and SMD) were degraded by less than 40%. The STZ removal efficiency was much higher in advanced process (27%) than that in conventional process (0%), while SMD showed the contrary result.
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As shown in Fig. 2(a) and Fig. 2(b), SMD was not reduced from flocculation to CL in the advanced process but had a removal efficiency of 63% during PCL→SF in the conventional
process.
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that
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enhanced
flocculation
(pre-/post-flocculation in conventional process, Fig. 1) should be an effective method to remove SMD. Overall, PO played a dominant role in SA removal during the advanced process for most SAs. Ozonation has been proved to be an effective method for SA removal
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[42]. Ozone could unselectively react with amino, aliphatic and active aromatic rings
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with the reaction rates of 1.0×105 to 2.5×106 M-1·s-1 at pH 7.0 [43,44]. After PO, FLO+SF seemed to have limited ability on the removal of most SAs, which could be
attributed to the highly mobile ability of SAs (low logKow) [45]. Compared with PO,
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OD contributed less to the SA removal, only 5 SAs (TMP, SAM, SDZ, SMZ, SMX)
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showed obvious removal, which was similar to a study in (about 50%) [41]. The low
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removal of SAs during PO might attribute to the low dosage of O3. SAs and other organic matter in raw water that were degraded into small molecules during PO and
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these small fractions would rapidly and competitively consume the low-level O3
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during OD. As for GAC, the adsorption capacity between the active carbon and trace
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organic compounds depends on hydrophobic interactions [46]. Adsorption through hydrophobic interactions tends to increase with an increasing logKow value of a substance [46]. Therefore, the low fractional removal of SAs in GAC can be
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attributed to SAs’ low logKow values (Table S1). Finally, the contribution of CL to SA removal in the advanced process was also quite limited, similar to that in the conventional process (Fig. 2(a)). 14
3.2 Occurrence and reduction of ARGs and intI1 in conventional and advanced DWTPs 3.2.1 Occurrence of ARGs and intI1 in raw water and finished water Recently, the detection of ARGs in drinking water has attracted much public
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attention [3]. Table 2 shows the occurrence of sul1, sul2 and class I integrase intI1 in
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raw water and finished water during the conventional and advanced processes. These three genes were detected in all water samples. Having the same water source, the
concentrations of sul1, sul2, intI1, and 16S rRNA in raw water of the two DWTPs
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(conventional and advanced processes) were similar. The relative concentrations of
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sul1, sul2 and intI1 in raw water ranges were 3.4×10-3 - 3.5×10-1, 3.7×10-4 - 3.4×10-2
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and 2.0×10-3 - 1.9×10-1 (ARGs/16S rRNA) (absolute concentration: 1.8×103 - 2.4×105 gene abundance/mL) (Table 2). The sul1 gene was more prevalent than sul2 and intI1
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genes in raw water. The concentrations of sul genes in our study were the same as that
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of Hangzhou Province (104 - 105 gene abundance/mL) [47], but higher than that in Michigan and Ohio (101 - 103 gene abundance/mL) [48]. The concentration of intI1 in
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our study was the same as that in Taihu Lake (103 - 104 gene abundance/mL) [49], but higher than that in Poland (102 gene abundance/mL) [50]. The differences among
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previous studies in other parts of the world and this work might be attributed to the high consumption of antibiotics in China compared to that in Europe and America (ten times higher) [2], which might pose a high selective pressure for environmental 15
microorganism. In addition to raw water, sul1, sul2 and intI1 were also largely detected in finished water of both DWTPs with the relative concentration of 9.7×10-5 - 9.9×10-1 (absolute concentration: 1.1×102 - 9.5×104 gene abundance/mL). The two processes
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showed a slight reduction in the levels of sul1, sul2 and intI1. Besides, although
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advanced drinking water treatment methods showed higher removal efficiencies for
SAs than conventional methods, conventional process had advantage in eliminating sul1, sul2 and intI1 (absolute concentration: conventional process: 4.64 log; advanced
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process: 3.96 log). Eventually the drinking water treated by the current water
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treatment processes (conventional or advanced process) would face a serious risk of
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ARGs pollution. Considering that ARGs are released from the antibiotic resistant bacteria (ARB) with different seasonal growth rates, the seasonal occurrence and
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reduction efficiency of ARGs in conventional and advanced treatment processes are
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discussed in Section 3.2.2 and Section 3.2.3.
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3.2.2 Reduction of ARGs and intI1 in conventional process Fig. 3 shows the log reduction and the relative concentration change of sul1, sul2,
and intI1 during the conventional process in four months, respectively. During PCL→
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SF in DWTP1, the absolute concentrations of all genes decreased (Fig.3 (a)), while the relative concentrations increased (except intI1 in July, September, and November) (Fig.3 (b)), which was consistent with the results of a study on ARGs removal in 16
Zhejiang province [47,51]. The increased of ARGs’ relative concentration might be due to that 16S rRNA was also affected by PCL→SF [25]. The suspended solids and waterborne microorganisms were trapped by the added coagulant/flocculant and removed by sand filtration [52]. Thus, this process may influence ARB that contains
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ARGs in their DNA, which could be confirmed by the reduction of 16S rRNA [53].
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However, the removal rates of sul1, sul2, intI1 and 16S rRNA during the PCL→SF
process varied in different months. The 16S rRNA achieved 1.5 log reduction in April, July, and November, but only 0.6 log reduction in September. The sul1 was reduced
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by more than 1.2 log concentration in April and November, while only by 0.6 and 0.4
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log in July and September. The sul2 gene only achieved a reduction of 1.0 log in
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November but less than 0.5 log reduction in other months. Among these three target genes, intI1 was the most easily removed one during the PCL→SF process, and it
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showed a higher reduction rate than sul1 and sul2 in July, September, and November
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(highest reduction of 1.9 log in November).
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The reduction efficiency was further weakened by CL. As shown in Fig. 3, some of the genes showed an increasing trend after CL. Especially in July and September, the absolute and relative concentrations of all three target genes (sul1, sul2 and intI1)
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increased after CL. A similar result of the ARG enrichment after chlorination was reported by Huang et al. [35]. The increase in its concentration after CL in all months could be attributed to the resistant to chlorine and antibiotics of microorganism [54]. 17
The class I integrons are reported to contain qacE△1 gene, which confers resistance to quaternary ammonium compounds, and sul1 gene [55]. The bacteria which harbors qacE△1 gene may survive during CL, thus, induce the increase of ARGs [56]. What’s more, the by-products of CL could promote the production of ARGs [57]. Meanwhile,
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the low dosage of free chlorine could help to increase the potential of ARG transfer to
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other bacterial cells [58]. Although some studies showed that CL could remove ARGs effectively [25,26], the chlorine dosage in those experiments was much higher (> 30 mg/L) than that in practical DWTPs (1.0 mg/L free chlorine in this study). At such
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low dosages, free chlorine can only destroy the cell wall of the microorganism
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including ARB; it cannot subsequently react with the released ARGs and integrase
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genes with double strands [59].
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3.2.3 Reduction of ARGs and intI1 in advanced process
Fig. 4 shows the monthly log reductions and relative concentrations of sul1, sul2,
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and intI1 during the advanced process. During PO, the reductions of sul1, sul2, and
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intI1 were unstable. The absolute concentrations of target genes (sul1, sul2, and intI1) increased after PO in all months except in November. The increased orders of
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magnitude of each gene were less than 0.6. However, the relative concentration of them decreased in all months except in July. As a strong oxidant, O3 can react with organic compounds in the cell wall and membrane and penetrate deeply into the cells 18
and inactivate the DNA [60,61]. Besides, several studies indicated that ozonation could inactivate ARGs at low O3 dosages [26,62]. However, due to the strong oxidizing ability of O3, organic matter (DOC: 2.3 mg/L in this study) in raw water, such as humic substances, carbohydrates, fatty acids as well as SAs, could consume
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O3 more quickly than ARGs [63], which explains the weak performance of ozonation
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in sul1, sul2 and intI1 reduction. The generation of bromate after ozonation could also
promote the abundance of ARGs [57]. With the decrease in precursors (including Br-) after the SF process and the increase in O3 dosage (PO: 0.5 - 1.0 mg/L; OD: 1.0 - 2.0
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mg/L), the reduction of target genes’ absolute concentration after OD showed a better
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performance than PO process (only sul1 increased in November), while the relative
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concentration increased in April and September. This result also explains the organic matter interference and O3 dosage shortage in PO. The FLO+SF process showed a
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similar phenomenon in the conventional process (Fig. 3); the reduction of sul1, sul2,
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and intI1 was positive in all months, except intI1 and 16S rRNA in November, and the
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relative concentration of three genes decreased except July. However, the log reduction of target genes after SF were much lower in the advanced process (Fig. 4(a)) than those in the conventional process (Fig. 3(a)) because of the enhanced
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flocculation (pre-/post-flocculation) in the conventional process (Table S2). Irregular reduction efficiencies of target genes were observed during GAC, and the relative concentration of them showed a falling trend in April and September but an ascending 19
trend in July and November (Fig. 4(b)), which might be related to the growth of different bacterial communities in different months and conditions [51]. Thus, GAC was not a stable technology for the removal of ARG and integrase genes. The results of the target genes reduction during PCL→SF (stable, Fig. 3),
IP T
PO/OD (unstable, Fig. 4), FLO + SF (stable, Fig. 4), and CL (unstable, Figs. 3 and 4)
SC R
show that the conventional drinking water treatment processes (flocculation, sedimentation, and sand filtration) play a more important role than the oxidation
processes (chlorination or ozonation) in practical DWTPs. Previous studies also
N
U
reported that the conventional DWTPS showed a more stable and positive reduction
M
A
in absolute concentrations of ARG than the advanced process [47,51].
Conclusion
ED
This study reported the occurrence and removal of 13 SAs, two sul ARGs, and
PT
class I integrase gene intI1 in two different DWTPS. The advanced treatment process
CC E
has a significant advantage over the conventional treatment method in the reduction of SAs, while showed an opposite performance in eliminating sul ARGs and intI1. The oxidation processes including pre-ozonation and post-chlorination could effectively
A
reduce SAs. However, those two processes re-increased the amount of sul1, sul2 and intI1 under certain conditions. The different monthly occurrences and reductions in sul1, sul2 and intI1 in each water treatment unit can help to optimize DWTPs’ 20
operations to manage the potential risks of ARGs in different months. Further investigation is needed to perform a health risk assessment of trace SAs in drinking water, the methods to eliminate them, and the potential transfer mechanisms of ARGs in specific drinking water treatment processes to reduce the effects of ARGs on
SC R
IP T
human health.
Acknowledgements
This work was supported by the National Water Pollution Control and
N
U
Management Technology Major Projects (No. 2017ZX07402003), Shanghai
A
Municipal Science and Technology Commission (No. 16DZ1204703) and Open
M
Project of State Key Laboratory of Urban Water Resource and Environment (No. QA201612), China Postdoctoral Science Foundation (No. 2017M621391), the
ED
Fundamental Research Funds for the Central Universities (No. 222201814055) and
A
CC E
PT
Shanghai Youth Science and Technology Sail Project (No. 18YF1406000).
21
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31
IP T
SC R
Fig.1 Schematic for the conventional and advanced drinking water treatment
A
CC E
PT
ED
M
A
N
U
processes.
32
(a)
1.2
Conventional process
PCL SF CL
0.8
0.6
IP T
Fractional removal of SAs
1.0
0.4
0.2
0 SG
SDZ SMR SMZ STZ SMT SMX SFX SCP SMD SAT
SC R
TMP SAM
(b) 1.2
Advanced process
U N
0.8
A
0.6
0.4
M
Fractional removal of SAs
1.0
PO FLO+SF OD GAC CL
ED
0.2
0
TMP SAM
SG
SDZ SMR SMZ STZ SMT SMX SFX SCP SMD SAT
PT
Fig. 2 Fractional removal of individual SAs by each drinking water treatment unit during (a) conventional
CC E
and (b) advanced processes (PCL→SF: pre-chlorination + pre-flocculation + rapid clarification +
post-flocculation + sand filtration; CL: chlorination; PO: pre-ozonation; FLO+SF: flocculation + sand
A
filtration; OD: post-ozonation; GAC: Granular activated carbon filtration). The red line means the total
removal efficiency was 80%, the height stands for the total remvoal efficiency of each SA
33
2.0
1.5
1.5
1.0
1.0
0.5
0.5
0.0
0.0
-0.5
-0.5 PCL
2.0
SF
CL
PCL 2.0
September
July
IP T
April
intI1
1.5
1.5
1.0
1.0
0.5
0.5
SF
November
SC R
sul2
CL
U
sul1
A
Log reduction of genes
2.0
16S rRNA
N
(a)
0.0
0.0
-0.5 SF
CL
A
CC E
PT
ED
PCL
M
-0.5
34
PCL
SF
CL
PCL
SF
FW 10
1
10
0
July
April 0
10
-1
10
-1
10
-2
10
-2
10
-3
10
-3
10
-4
10
-4
10
1
sul1
sul2
intI1 10
September 10
sul1
1
November
0
10
0
-1
10
-1
10
-2
10
-2
10
-3
10
-4
sul2
intI1
N
10
sul1
-3
A
10
sul2
10
intI1
-4
sul1
sul2
intI1
ED
M
Relative concentration of ARGs
10
IP T
RW
(b)
SC R
1
U
10
Fig. 3 Log reduction of target genes by conventional process (PCL→SF:
filtration;
CL:
chlorination;
RW:
CC E
sand
PT
pre-chlorination + pre-flocculation + rapid clarification + post-flocculation + Raw
water;
A
water).(a)absolute concentration;(b)relative concentration
35
FW:
Finished
sul2
intI1
April
1.0
1.0
0.5
0.5
0.0
0.0
-0.5
-0.5
-1.0
-1.0
-1.5
-1.5 PO
1.5
July
1.5
FLO+SF
OD
GAC
PO
CL
September
1.5 1.0
0.5
0.5
0.0
0.0
-0.5
-0.5
OD
GAC
CL
GAC
CL
November
N
U
1.0
FLO+SF
IP T
Log reduction of genes
1.5
sul1
16S rRNA
SC R
(a)
-1.0
A
-1.0
-1.5
FLO+SF
OD
GAC
CL
A
CC E
PT
ED
PO
M
-1.5
36
PO
FLO+SF
OD
RW
PO
FLO+SF
OD 10
April
1
0
10
0
-1
10
-1
10
-2
10
-2
1
sul1
sul2
intI1 10
sul1
0
September
November 10
-1
10
-1
10
-2
10
-2
10
-3
10
-3
sul2
intI1
N
U
0
10
FW
July
10
10
GAC
IP T
(b)
10
sul1
A
Relative concentration of ARGs
10
1
SC R
10
sul2
intI1
sul1
sul2
intI1
A
M
A
-4
ED
Fig. 4 Log reduction of target genes by advanced process (PO: pre-ozonation;
PT
FLO+SF: flocculation + sand filtration; OD: post-ozonation; GAC: Granular activated carbon; CL: chlorination; RW: Raw water; FW: Finished
A
CC E
water).(a)absolute concentration;(b)relative concentration
37
I N U SC R
Table 1-Occurrence of sulfonamides in raw water and finished water of conventional and advanced processes.
Raw water (ng/L)
iotics
ax.a
TMP
in.b
CC E
0.86
M
.72
3
F
ean.c
req.d
1
1
1
PT
4
M
ED
M
Finished water (ng/L)
M
Antib
Advanced process
A
Conventional process
M ax.
M in.
M ean.
2
F req.
0
Finished water (ng·L)
Raw water (ng/L) M ax. 6
M in.
3
M ean.
1
F req.
3
M ax.
1
M in.
0
00%
1.
.14
1
F req.
0
2
0 .76
1
ean.
0
0 0.60
M
0
7% 0
.79 1
.02 4
.10 1
00% 2
.89
1
.14 1
0
9% 0
1
SAM
A
.16
.88
75
0
0.
.08 0
.19
.80
00%
0
1
0
.32
0
00% 0
0
.07 3
38
0
.05
.75
00% 0
0 .01
5
.31
1
0 %
7
.09
0
4% 1.
.21
0 0
01 4
.30
1
0
SG
SDZ
00%
0
0
4% 1
% 7
0
0
0
2
I 68 0
1% 0.
0
3% 4.
0.81
STZ
.87
10
.72
CC E
12
3%
3% 1
.28
0
7 1%
2
.96 4
.23 0
0 .59
1 00%
0
.45
.50 1
0
3%
0
.36
1.97
0
0.29
3 3%
0
.69 1
0
0.
00% 2
0
4
.21 1
.06 0
.64
.34
0
00%
4 3%
0
0
.39
4
.07 0
2
.36 3
.15 0
3%
2
3%
0
7
.38 6
4
2.05
7
.11 0
0
3% 0
4
1% 1
.22 1
.08 1
3% 2
4
0 .23
0
4
0 .21
5
0
.31
0 .12
9%
0
0 7.27
9%
0
0 .12
SMX
SFX
0
0 18
4
0 .05
0
6
1
1%
0
4
0.
.33
3
.73
0
0
SMT
.14
2
0
.29
A
00%
0.
PT
0
ED
SMZ
.71
0
.45
1
3%
0
A
17 1
.44
0
M
.51
.43 4
0
SMR
N U SC R
.16
7% 0
5.44 0
39
.74 0
9.85 0
0
00% 1
1.24
.53 0
0
3% 0
0
I 13 1
0
9% 1.
29
2
0.
ED
.14
75
1%
CC E -
.73
3
.10 6
0
.21
0
3% 0
.88 2
9%
0
3 3%
0
2.92
-
0 0
.11 0
.78 3
3%
0 %
4
0
0
0
3% 0
.55 1
.19 0
1.41
4%
-
.16
.02
A
(a Maximum concentrations; b Minimum concentrations; c Mean concentrations; d Frequency of detection)
9% 4
4.73
40
2
0 .06
-
3% 0
3
-
4
0 .25
.44
0
6% 0
1 -
8
0 .05
4
%
0 .27
.39
2
.85
0 18
4%
0
0
0
.23
0
.71
7
0.
PT
1
Total
00%
0
SMD
SAT
A
.77
0
M
.82
%
1
SCP
N U SC R
.48
-
-
.19
I N U SC R
Table 2 - Occurrence of ARGs in raw water and finished water of conventional and advanced processes.
A
Conventional process Raw water
M
Genes
sul1
PT
concentrationc Relative
sul1
CC E
concentrationd Absolute
Finished water
Raw water
Finished water
Max.a
Min.b
Max.
Min.
Max.
Min.
Max.
Min.
2.1×10
1.3×10
6.0×10
9.8×10
2.4×10
1.1×10
9.5×10
1.1×10
ED
Absolute
Advanced process
5
4
4
2
5
4
4
2
1.3×10
3.4×10
9.9×10
9.5×10
3.5×10
4.1×10
9.7×10
9.7×10
-1
-3
-1
-3
-1
-2
-1
-5
1.6×10
1.8×10
9.9×10
1.3×10
9.8×10
2.2×10
6.2×10
1.3×10
sul2
A
concentration Relative
4
3
3
2
3
3
3
3
3.4×10
3.7×10
4.2×10
1.2×10
1.5×10
8.0×10
1.3×10
2.0×10
-2
-4
-1
-3
-2
-3
-1
-3
8.5×10
1.1×10
5.2×10
5.0×10
1.5×10
2.3×10
7.6×10
1.7×10
sul2 concentration intI1
Absolute
41
I Relative
4
4
1.9×10
rRNA
concentration
M
Absolute
-1
5.3×10 6
ED
16S
2.0×10
4
2
4
3
3
2
8.5×10
4.8×10
2.0×10
7.3×10
2.8×10
1.6×10
-3
-1
-3
-2
-3
-1
-4
4.5×10
1.7×10
2.1×10
3.0×10
3.1×10
1.1×10
2.1×10
A
intI1 concentration
N U SC R
concentration
5
5
4
6
A
CC E
PT
(a Maximum concentrations; b Minimum concentrations; c copy abundance/mL; d ARGs/16S rRNA)
42
5
6
4
