Contributions of volatilization, photolysis, and biodegradation to N‑nitrosodimethylamine removal in conventional drinking water treatment plants

Science of the Total Environment 697 (2019) 133993

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

Contributions of volatilization, photolysis, and biodegradation to N nitrosodimethylamine removal in conventional drinking water treatment plants Yu Qiu a, Er Bei a, Shuguang Xie b, Shixiang Li a, Jun Wang a,c, Xiaojian Zhang a,c, Stuart Krasner d, Chao Chen a,c,⁎ a

State Key Joint Laboratory of Environmental Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China State Key Joint Laboratory of Environmental Simulation and Pollution Control, School of Environment, Peking University, Beijing 100871, China c Research Institute for Environmental Innovation (Suzhou), Tsinghua, Suzhou 215163, China d La Verne, CA 91750, USA b

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• NDMA could be degraded during water treatment processes. • Photolysis plays the most important role in removing NDMA in conventional DWTPs. • NDMA removal in biotreatment can't be ignored and needs further investigation. • Volatilization presented negligible NDMA removal.

a r t i c l e

i n f o

Article history: Received 2 April 2019 Received in revised form 12 August 2019 Accepted 18 August 2019 Available online 20 August 2019 Editor: Ching-Hua Huang Keywords: N nitrosodimethylamine Volatilization Photolysis Biodegradation Drinking water treatment plants

a b s t r a c t N nitrosodimethylamine (NDMA) was detected in the source water of some Chinese drinking water treatment plants (DWTPs), which decreased in concentration along the treatment train. Volatilization, photolysis, and/or biodegradation were suspected of being capable of attenuating NDMA. In this study, the contribution of these mechanisms to NDMA removal was investigated by a field study in a conventional DWTP with aerated biopretreatment, as well as in laboratory-based experiments. The effluent of each unit process (i.e., aerated biopretreatment tank, horizontal sedimentation tank, sand filter) of this DWTP was sampled in the winter and summer, and the concentration of NDMA, its formation potential, and other water quality parameters were measured. NDMA removal by volatilization and biodegradation was simulated in batch experiments, and that by photolysis was calculated with parameters reported in the literature. The sampling results indicated that the aerated biofilm reactor of this DWTP removed 48% of the NDMA in August and 22% in December. According to modeling results, it could be well explained by photolysis (NDMA removal of 51% in summer and 25% in winter) and biotreatment (NDMA removal of 0.2–12% in summer and 0.1–6.1% in winter), with little contribution from aeration (NDMA removal of 0.8%). The sampling results indicated that the sedimentation tank removed 19% of NDMA in August and 9.2% in December. According to modeling results, it could be well explained by photolysis (NDMA removal of 16% in August and 9.4% in December), but little by volatilization. Thus, photolysis was shown

⁎ Corresponding author at: State Key Joint Laboratory of Environmental Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China. E-mail address: [email protected] (C. Chen).

https://doi.org/10.1016/j.scitotenv.2019.133993 0048-9697/© 2019 Elsevier B.V. All rights reserved.

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Y. Qiu et al. / Science of the Total Environment 697 (2019) 133993

to be the most important process for NDMA removal in this DWTP. Further investigation is needed to better understand NDMA removal during biotreatment. © 2019 Elsevier B.V. All rights reserved.

1. Introduction N nitrosamines, especially N nitrosodimethylamine (NDMA), have received much attention from the water industry recently because they are strongly carcinogenic and are frequently found in chloraminated drinking water (Choi and Valentine, 2002; Mitch and Sedlak, 2002; Krasner et al., 2013). Preoxidation with ozone or chlorine can destroy NDMA precursors (Krasner et al., 2013). However, there are reports about NDMA formation after ozonation (Schmidt and Brauch, 2008) and one after prechlorination (Chen et al., 2016). The US Environmental Protection Agency's (USEPA's) Integrated Risk Information System database indicates that NDMA at a concentration of 0.7 ng/L is associated with a 10−6 lifetime excess cancer risk (US EPA, 1997). NDMA was included in the USEPA's Unregulated Contaminant Monitoring Rule 2 (UCMR2) (USEPA, 2007) and Contaminant Candidate List 3 (USEPA, 2009). The California State Water Resources Control Board notification levels for NDMA, N nitrosodiethylamine, and N nitrosodipropylamine are all 10 ng/L (CSWRCB, 2016). There is currently no national regulation for NDMA, but there is serious concern about NDMA in China (Bei et al., 2019). In a survey of China, the occurrence of N nitrosamines in 164 water samples (including source water, finished water, and tap water) were measured and compared to in the USA (according to the UCMR2 dataset) (Russell et al., 2012; Bei et al., 2016b). The concentrations of NDMA in the detected samples showed higher levels in China than in the USA. The average and the median in China was 33 and 22 ng/L, respectively, whereas those in the USA were 9 and 4 ng/L, respectively. In June 2018, Shanghai was the first Chinese city to set a local NDMA limit, 100 ng/L. The nitrosamine issue in China could be partially attributed to the anthropogenic pollution of the water sources (Bei et al., 2016a). It has been reported that NDMA may be present in industrial discharges, such as rubber manufacturing, leather tanning, pesticide manufacturing, food processing, foundries, and dye manufacturing; as a result, NDMA can be present in sewage treatment plant effluents (Liteplo and Meek, 2002). Industrial or domestic wastewater discharge have been shown to be significant sources of nitrosamine precursors (Sedlak et al., 2005). Therefore, drinking water treatment plants (DWTPs) with source waters impacted by industrial or wastewater discharges may have unacceptable concentrations of NDMA (Mitch et al., 2003). Reverse osmosis removed 24–56% of NDMA at recycling plants (Plumlee et al., 2008). Ultraviolet (UV) irradiation at a high fluence with H2O2 present (an advanced oxidation process) is the most accepted method for removing NDMA in water-recycling facilities. NDMA removal efficiencies of 43–66% have been found for the UVadvanced oxidation process (Plumlee et al., 2008). Although UV irradiation effectively removes NDMA, the use of UV irradiation in DWTPs is limited by the high UV fluence (approximately an order of magnitude higher than normally used to inactivate pathogens) required to give an acceptable NDMA removal efficiency (Sedlak et al., 2005). Moreover, at DWTPs that add chloramines at the end of the plant, most of the NDMA is formed in the distribution system. Thus, application of UV in the plant will not control the bulk of the NDMA formed. Therefore, for DWTPs with NDMA in the source water (as has been found in China), attention at the DWTP has focused on natural attenuation processes (i.e., volatilization, photolysis, and biodegradation). Weak absorbance at 330 nm (n → π ∗ transition) means NDMA photolysis can occur in natural sunlight (Stefan and Bolton, 2002). Plumlee and Reinhard (2007) found an NDMA half-life of 16 min under direct

photolysis at 765 W/m2, which represented natural irradiation at noon in midsummer in Southern California. Chen et al. (2010) performed outdoor solar photolysis tests and found an NDMA half-life of 8.5 min at irradiation intensities of 1150–1300 W/m2. Volatilization may also remove NDMA. In laboratory experiments, Haruta et al. (2011) found NDMA has a dimensionless Henry's law constant of 1.0 × 10−4, suggesting that NDMA has a relatively strong potential to volatilize and that volatilization may be an important NDMA transport pathway. Williams et al. (2014) found that 20% of the NDMA in recycled water in a natural reservoir system volatilized in 14 d. Jin et al. (2012) concluded that, in natural watersheds, NDMA photolysis rate occurs 0.8–300 times that of NDMA volatilization, and volatilization would be accelerated in stream-type watersheds with high water velocities and shallow water. The volatilization rates could reach 10−4– 10−3 h−1. Biodegradation of NDMA has recently received attention. The dimethylamine group in NDMA is strongly electron donating, so NDMA is readily biodegraded (Wijekoon et al., 2013). Wijekoon et al. (2013) found 94% of the NDMA present was removed in an aerobic membrane bioreactor (MBR). Studies of NDMA biodegradation in an aquifer and soil and by an isolated microbial strain have been performed (Gan et al., 2006; Szecsody et al., 2008; Patterson et al., 2012; Wang et al., 2015b). Wang et al. (2015b) found that Rhodococcus cercidiphylli A41 AS-1 was a nitrosamine-reducing bacterium in a biological activated carbon filter, and the isolated strain could remove 85% of NDMA. The half-life of NDMA biodegradation was reported to be 12–35 days under aerobic conditions and 26–38 days under anaerobic conditions in one investigation (Gunnison et al., 2000). More than 90% of Chinese DWTPs use conventional treatment processes, i.e., coagulation, sedimentation, filtration, and disinfection (Bei et al., 2019). An important strategy for controlling NDMA formation in drinking water is to remove NDMA precursors before disinfection (Krasner et al., 2013). Alternatively, NDMA itself is removed at advanced wastewater treatment facilities (Plumlee et al., 2008). Little research on direct NDMA removal in full-scale DWTPs has been performed. The publications mentioned above indicate that natural attenuation (by photolysis, volatilization, and biodegradation) may contribute to NDMA removal. However, none of these studies quantified the contribution of these mechanisms. Therefore, NDMA removal by these mechanisms should be quantified when developing comprehensive, precise, and economical strategies for controlling NDMA and other nitrosamines. This study had three objectives: (1) to monitor the removal of existing NDMA in full-scale conventional DWTPs and at a DWTP with bio-pretreatment, and to evaluate the NDMA removal performance of each treatment process at the latter plant, (2) to develop NDMA photolysis, volatilization, and biodegradation models and to quantify NDMA removal in each treatment process, and (3) to compare actual and calculated NDMA removal efficiencies and determine the contribution of each mechanism to the total NDMA removal efficiency. As far as the authors know, this is the first comprehensive and quantitative study on the attenuation of source-water derived NDMA in a DWTP by photolysis, biodegradation, and volatilization, although the characteristics of photolysis (Plumlee and Reinhard, 2007; Afzal et al., 2016), biodegradation (Wang et al., 2015b; Wang et al., 2015c), and volatilization (Haruta et al., 2011) have been investigated individually before in other fields. Preoxidation (with ozone or chlorine) has been shown to destroy NDMA precursors (Krasner et al., 2013), but it has not been shown to destroy NDMA itself. Alternatively, ozone (Schmidt

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and Brauch, 2008) and in one case prechlorination (Chen et al., 2016) has been shown to form NDMA.

concentration and microbial community structure in each sample were determined.

2. Materials and methods

2.4. NDMA and FP analysis

2.1. Water sample collection

The NDMA concentrations in the samples collected during the laboratory experiments were in microgram/l levels, so no pretreatment (extraction/concentration) was necessary. A 1 L aliquot of a DWTP sample was extracted using a Dionex Auto Trace 280 solid-phase extraction (SPE) instrument (Thermo Fisher Scientific, Waltham, MA, USA) fitted with a coconut charcoal cartridge (CNW Technologies, Shanghai, China), and the cartridge was eluted to give a 1 mL extract. Each extract was then analyzed by liquid chromatography/tandem triple quadrupole mass spectrometry using an Agilent 1260 system coupled to an Agilent 6460 system (Agilent Technologies, Santa Clara, CA, USA). Separation was achieved using a C8 BEH column (2.1 mm internal diameter, 100 mm long, 1.7 μm particle size; Waters, Milford, MA, USA) and a mixture of acetonitrile and 0.1% formic acid solution as the mobile phase. Information on the reagents used are provided in S1. The liquid chromatography/tandem mass spectrometry parameters are shown in S3 (Tables S.1 and S.2). NDMA FP analysis was conducted for source water, biofilm tank effluent, and sedimentation effluent. The procedure used to determine the NDMA FP was similar to the procedure used in a previous study (Liao et al., 2014). Excess monochloramine (40 mg/L as Cl2) was added to a 500 mL aliquot of a DWTP sample. The solution was adjusted to pH 8 by adding 0.02 mol/L of phosphate buffer, then the bottle was stored in the dark at 25 °C for 3 d. The reaction was then quenched by adding excess sodium thiosulfate (2 mL, 1 mol/L), then the solution was analyzed for NDMA as described above. The final FP value was defined as the NDMA concentration after the FP test minus the original NDMA concentration.

Samples of the source and finished water (after chlorination) were collected from five conventional DWTPs (labeled DWTP1–DWTP5) in Southeast China that use a free-chlorine disinfection process and from one conventional DWTP (DWTP6) in Central China that uses chlorine dioxide and chlorine in the disinfection process. Samples of effluent from each unit process at a conventional DWTP (DWTP7) with biopretreatment in East China were also collected (i.e., source water, biofilm tank, sedimentation tank, sand filter). The sampling campaigns at DWTP7 were conducted twice in one year, in August (average water temperature was 32.4 °C) and December (average water temperature was 12.0 °C). DWTPs 1–5 were sampled once in July. DWTP6 was sampled quarterly in one year. DWTP7 has a capacity of 50,000 m3/d and uses an aerated biofilm pretreatment tank (this process is widely applied at DWTPs in China that have high concentrations of ammonia in their source waters) and then conventional treatment (i.e., coagulation, sedimentation, sand filtration, and free chlorination) (as shown in Fig. 1). The parameters for the main processes at DWTP7 are shown in Table 1. A sample of 3–4 L was collected from each sampling site and stored in a 4 L amber glass bottle. Each finished water sample was dechlorinated immediately by adding an excess amount of sodium thiosulfate. The samples were transported in an ice box to Tsinghua University, then each sample was passed through a Millipore 0.45 μm membrane filter (Merck, Darmstadt, Germany) and then stored at 4 °C until analysis. The NDMA concentration, NDMA formation potential (FP), and other water quality parameters were determined within 7 d of sample collection. 2.2. Batch volatilization experiments Volatilization experiments were performed using an aeration reactor wrapped in aluminum foil (schematic of the aeration apparatus is shown in Fig. S.1, and more details are provided in the Supplementary Material [Section S2]). A 1.5 L aliquot of a 50 μg/L NDMA solution was transferred to the glass vessel at room temperature (22 °C) and then aerated using a gas flow rate of 25, 60 or 200 mL/min. Samples were collected after 0, 10, 20, 30, 60, 120, 240, 360, and 480 min, and the NDMA concentrations in the samples were determined. A control experiment was performed using the reactor without aeration (with the air compressor turned off and the solution stirred at 300 r/min with a magnetic stirring bar) using an initial NDMA concentration of 100 μg/L. Samples were collected after 0, 0.5, 1, 2, 4, 6, 8, 24, 48, and 69 h. The NDMA concentrations in the samples were determined. 2.3. Batch biodegradation experiments A water sample was collected from a natural water body with a native microbial community. The total bacterial count was 2100 CFU/mL. NDMA was added to an aliquot of the natural water sample to give an NDMA concentration of 20 μg/L, and the solution was aerated in the dark in an incubator at 30 °C. A control solution prepared using ultrapure water was kept under the same conditions. Samples of the experimental solution were collected after 0, 1, 2, 3, 5, and 7 days, and samples of the control solution were collected after 0, 3, and 7 days. The time points were chosen according to the half-life of NDMA during biodegradation as reported before (Gunnison et al., 2000). The NDMA

2.5. Water quality analysis Dissolved organic carbon (DOC) and total dissolved nitrogen (TN) were determined using a Shimadzu TOC-VCPH analyzer and a TNM-1 system, respectively (Shimadzu, Kyoto, Japan). The NH3-N concentration was determined using a spectrophotometric method using Nessler's reagent. The nitrate and nitrite concentrations were determined using a Dionex ICS-5000 ion chromatograph (Thermo Fisher Scientific, USA). A UV-2700 UV–Vis spectrophotometer (Shimadzu, Kyoto, Japan) was used to acquire the UV–visible spectrum of the source water sample and absorption at a wavelength of 254 nm for all of the samples. Excitation–emission matrix (EEM) fluorescence spectra were acquired using an F-7000 fluorescence spectrometer (Hitachi, HighTechnologies, Tokyo, Japan). The fluorescence spectrometry procedure is described in more detail in the Supporting Materials (S4, Table S.3). These water quality parameters did not show any clear relationship with NDMA FP by Pearson analysis. 2.6. Microbial community analysis Microbial community analysis was performed at Peking University. Each water sample was passed through a 0.22 μm Millipore membrane filter (Merck). The filter was stored at −20 °C until analysis (Zhang et al., 2019). DNA was extracted from the membrane using sodium dodecyl sulfate, and the DNA obtained was examined by agarose gel electrophoresis using a gel concentration of 1%. Polymerase chain reaction (PCR) amplicon libraries were constructed using the bacterial primers 515F (5′-GTGCCAGCMGCCGCGG-3′) and 806R (5′-GGAC TACHVGGGTWTCTAAT-3′) targeting V4 hypervariable regions (Caporaso et al., 2012). The sample was mixed and the PCR products were purified, then a gene library was constructed using a TruSeq DNA PCR-Free Sample Preparation Kit (Illumina, San Diego, CA, USA). Microbial community analysis was then performed using the HiSeq

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Y. Qiu et al. / Science of the Total Environment 697 (2019) 133993

(a) Process flowchart of DWTP7 (“*” represents the sampling points)

(b) Biofilm tank for ammonia-nitrogen removal

(c) Coagulation-sedimentation process

(d) Sand filtration process Fig. 1. Process flowchart and photos of DWTP7.

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Table 1 Parameters for the main processes at DWTP7.

Aerated biofilm tank Sedimentation tank Sand filter

Structure capacity (m3/d) × number

Structure size (length × width × height, m)

Water depth (m)

Hydraulic retention time (h)

Other parameters

Covered

12,500 × 4 50,000 × 1 6250 × 8

5.7 × 6.4 × 6.5 89.3 × 11.8 × 3.3 3.4 × 12.4× 4.45

6 3 4

1 1.5 0.5

Air-water ratio = 1 – Filtration rate = 8.3 m/h

No No Yes

2500 platform (Novo Gene Biological Information Technology, Beijing, China).

that the source water might have been impacted by wastewater discharges (Barker and Stuckey, 1999).

3.1.2. Removal of NOM in the 7 DWTPs The other water quality parameter results are shown in S5 (Table S.4 and Fig. S.2). This includes the occurrence and removal of DOC and EEMs, both indicators of NOM. According to Fig. S.2, the source water in DWTP7 had major EEM peaks in Region II, which was related to aromatic proteins II, and Region IV, which was related to soluble microbial products. The presence of soluble microbial products is an indication

3.1.3. Performances of each unit treatment process in DWTP7 Source water, aerated biofilm tank effluent, and conventional treatment process effluent samples (before the disinfection process) from DWTP7 were analyzed to allow the performance of each treatment process to be assessed. The dynamics of NDMA and NDMA FP along the treatment train are shown in Fig. 4. As shown in Fig. 4(a), the NDMA concentrations in the source water in August and December were 22 and 17 ng/L, respectively. The biofilm pretreatment tank removed 48% of the NDMA in August and 22% in December, indicating that the temperature or some other seasonal parameters may have affected NDMA removal in the biofilm tank. Less NDMA was removed by the conventional treatment processes (which followed after the biofilm tank) (19% in August and 9.2% in December) than by the biofilm pretreatment. As shown in Fig. 4(b), the NDMA FPs in the source water in August and December were 39 and 98 ng/L, respectively. The aerated biofilm and conventional treatment processes removed 21 and 12%, respectively, of the NDMA FP in August and 26 and 20%, respectively, of the NDMA FP in December. Conventional treatment processes have been found to remove only limited amounts of NDMA FP in previous laboratory experiments (Liao et al., 2015a; Liao et al., 2015b), pilot-plant tests (Liao et al., 2014), and full-scale DWTP studies (Wang et al., 2013; Bei et al., 2016b). The removal of NDMA in an aerated biofilm pretreatment tank could be attributed to biodegradation, photolysis, and/or volatilization. Photolysis and volatilization could also have occurred in the open horizontal sedimentation tank with a water depth of 3 m, which has a large surface area and a long hydraulic retention time. The medium in the sand filter could provide a niche for microbes capable of biodegrading NDMA (Wang et al., 2015b). However, the empty bed contact time of this filter is about 15 min, which is too short for NDMA to be biodegraded greatly. Jin et al. (2012) found that insignificant amounts of NDMA (b1% of the

Fig. 2. N nitrosodimethylamine (NDMA) concentrations in the source and finished water samples from the five water treatment plants (DWTPs) using conventional processes (the samples were collected in July 2016; each error bar represents the standard deviation for triplicate measurements).

Fig. 3. N nitrosodimethylamine (NDMA) concentration in the DWTP6 influent and effluent (each error bar represents the standard deviation for triplicate measurements).

3. Results and discussion 3.1. NDMA dynamics through the treatment processes 3.1.1. Removal of NDMA in the conventional DWTPs Five DWTPs in Southeast China using conventional treatment processes were studied. The NDMA concentrations for the source water and finished water samples are shown in Fig. 2. All five source water samples contained NDMA with a range of 40 to 85 ng/L and a mean concentration of 55 ng/L. NDMA may have been present in the source water because of upstream wastewater discharges, as reported in some Chinese DWTPs (Wang et al., 2016). In DWTPs, no more NDMA would be produced before the disinfection process, where chloramines preferentially form NDMA (and chlorine typically does not). In the finished water of the five DWTPs, the concentration of NDMA decreased greatly, with a mean removal of 45%. As shown in Fig. 3, a decrease in NDMA was also observed in DWTP6. Note, NDMA in this plant influent was sampled after preoxidation with chlorine dioxide and chlorine in the presence of ammonia (thus, chloramines may have formed). Therefore, NDMA may have present in the source water and/or was formed during the preoxidation step. These results suggested that NDMA could be removed within conventional DWTPs.

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Fig. 4. (a) N nitrosodimethylamine (NDMA) concentrations and (b) NDMA formation potentials (FPs) in DWTP7 (each error bar represents the triplicate measurement range).

total NDMA concentration) become adsorbed to suspended particles in surface water because NDMA has a low octanol–water partition coefficient. Sand filters in DWTPs, therefore, probably remove negligible amounts of NDMA. In January 2018, we found the NDMA concentrations in the sedimentation tank and sand filter effluents were almost the same, 12 and 13 ng/L, respectively. In a previous monthly study of another DWTP, we also found that the sand filter removed only negligible amounts of NDMA (Wang et al., 2015a). Thus, further research on sand filtration was not conducted. The contributions of photolysis, biodegradation, and volatilization to NDMA removal in the aerated biofilm tank and sedimentation tank were focused on in the following experiments. 3.2. NDMA removal by biodegradation 3.2.1. Removal of NDMA in a natural water The biodegradation of NDMA was investigated in batch experiments. The results are shown in Fig. 5. The degradation process can be regarded as a pseudo-first-order reaction (Patterson et al., 2012) that can be described by Eq. (1), C ¼ C0 e−KB t

ð1Þ

where C is the concentration (μg/L) at time t (d), C0 is the initial concentration (μg/L) and KB is the biodegradation rate (d−1).

Fig. 5. Effects of biodegradation on the N nitrosodimethylamine (NDMA) concentrations in batch experiments.

The calculated NDMA aerobic degradation rate in batch experiment was 0.05 d−1 and the half-life was 13.9 d based on the data in Fig. 5. 3.2.2. Microbial community The natural water did not contain R. cercidiphylli A41 AS-1, which was reported for NDMA biodegradation in a previous study (Wang et al., 2015b). However, the aerobic denitrifiers Achromobacter, Acinetobacter, and Pseudomonas were detected. These all belong to the Proteobacteria phylum and are believed to be favored under oligotrophic conditions (Zhu et al., 2012). The low ammonia concentrations during the sampling campaigns (as shown in Table S.4), may have caused nitrosamines to be biodegraded through being used as nitrogen sources by these microbes. It has previously been found that Pseudomonas can degrade N nitrosamines (Nakahama and Urakubo, 1987). The relative abundances of the Achromobacter, Acinetobacter, and Pseudomonas bacteria are shown in S6 (Table S.5) and Fig. 6. These results largely agreed with the results of previous studies in which the Proteobacteria phylum (mostly α-, β-, and γ-proteobacteria) were found to be the dominant bacteria in drinking water biofilters (Liao et al., 2013a; Liao et al., 2013b; Wang et al., 2015a). 3.2.3. Microbial performance assessment Besides the NDMA biodegradation rate obtained in our batch experiments, Wijekoon et al. (2013) reported that an aerobic MBR removed 94% of the NDMA present (in 5.0 ± 0.5 g/L mixed liquor suspended solids) at 30 ± 0.1 °C in 24 h. The NDMA biodegradation rate in the MBR was calculated to be 2.8 d−1 using Eq. (1). The bacteria counts in the biofilm tank (1 mg/m3 mixed liquor volatile suspended solids) were generally between what was in the natural water (2100 CFU/mL) and that of MBRs (5.0 ± 0.5 g/L mixed liquor suspended solids) (Wijekoon et al., 2013). Thus, the NDMA

Fig. 6. Microbial community structures (at the phylum level) in the natural water that was analyzed.

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biodegradation rate in the biofilm tank was estimated to be between 0.05 and 2.8 d−1 at 30 °C. The Arrhenius equation is KT ¼ K20 θT ðT−20Þ

ð2Þ

where θT is the temperature coefficient (1.035), KT is the biodegradation rate at T °C, and K20 is the biodegradation rate at 20 °C. The K20 value for the biofilm tank, calculated using the Arrhenius equation, was between 0.04 and 2.0 d−1. The calculated biodegradation rates for August (at 32 °C) and December (at 12 °C) were 0.05–3.0 and 0.03–1.5 d−1, respectively. According to Eq. (1), the NDMA removal efficiencies for the biofilm tank in August and December were 0.2–12% and 0.1–6.1%, respectively. 3.3. NDMA removal through aeration or volatilization

7

where C1 is the concentration at the outer edge of the aqueous film, C10 is the initial concentration in the water body, KL is the overall mass transfer coefficient for the liquid phase (cm/h), and Z is the mean depth of the water body (cm). And 0

KH ¼

1 kg ð1=K L −1=kl Þ

ð6Þ

where KL is the overall mass transfer coefficient for the liquid phase 0

(cm/h), K H is Henry's law constant, kg is the exchange coefficient in the gas phase (cm/h), and kl is the exchange coefficient in the aqueous phase (cm/h). According to Eq. (4), the KL/Z value was 0.0013 h−1, and the mean water depth was 31 cm. Then, KL was calculated to be 0.041 cm/h. Haruta et al. (2011) reported the exchange coefficients kg and kl to be 0

3.3.1. NDMA removal through aeration Aeration removed some NDMA, as shown in Fig. 7(a). The curves for the three aeration flow rates were very similar, and the small differences between them may have been caused by the bubbles being slightly larger at higher aeration flow rates than at lower aeration flow rates. These results confirmed that NDMA removal through aeration was related to the gas–water ratio rather than aeration flow rate. The overall mass transfer coefficient (G) was, therefore, able to be calculated. The mean overall mass transfer coefficient was determined to be 0.0090, 0.0084, 0.0062 min−1 respectively in three tests with different fluxes (0.0079 min−1 on average). Then, the relationship between the removal efficiency and gas–water ratio was described as Eq. (3). C 1− ¼ 1−e−0:0079q C0

ð3Þ

where C0 is the initial concentration (ng/L) and q is the gas–water ratio. 3.3.2. NDMA volatilization rate without aeration The change of NDMA concentration in the non-aerated solution was shown in Fig. 7(b). According to the experimental results, the NDMA dynamic followed the pseudo first-order reaction and could be fitted as Eq. (4). C ¼ e−0:0013t C0

ð4Þ

According to the two boundary-layer theory, the following equations could be applied. Cl ¼ Cl0 e−KL t=Z

ð5Þ

300 and 2 cm/h, respectively. The NDMA K H value was calculated to be 1.38 × 10−4 using Eq. (6), agreeing with the value of 1.0 × 10−4 determined by Haruta et al. (2011). Then according to Eq. (5), for a specified water depth, the removal of NDMA can be calculated using Eq. (7), 1−

C ¼ 1−e−0:041t=Z C0

ð7Þ

where C0 is the initial concentration (ng/L), t is time (h), and Z is the mean water depth (cm). 3.4. Simulation of NDMA removal through photolysis Weak absorbance by NDMA at 330 nm means NDMA can absorb and be degraded by natural sunlight. Biofilm and sedimentation tanks in DWTPs have long retention times and large surface areas, so direct photolysis of NDMA by natural sunlight could occur in these systems. 3.4.1. Building a photolysis model Sunlight attenuation with increasing depth was a key factor in the direct photolysis model. Sunlight attenuation was divided into two parts, absorption by NDMA during direct photolysis and absorbance by other organic substances. The direct photolysis model was built using the Schwarzenbach (2015) direct photolysis theory equation. h i ka ðλÞ ¼ WðλÞεðλÞ 1−10−αD ðλÞ Z ZαðλÞ

ð8Þ

where W(λ) is the spectral photon fluence rate (Einstein·cm−2·s−1), ε(λ) is the molar extinction coefficient for NDMA (M−1 m−1), αD(λ) is the apparent attenuation coefficient (m−1) (αD(λ) = D(λ)α(λ), where D(λ) is the distribution function for light scattering in the

Fig. 7. Removal of N nitrosodimethylamine during the aeration experiments at (a) different flow rates and (b) without aeration (but with stirring at 300 r/min).

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water, approximately 1.2 for non-turbid shallow water), Z is the water depth (m), and ka(λ) is 10 Einstein·s−1·(mol NDMA)−1. Eq. (9) is used for direct photolysis in Milli-Q water.

ka ðλÞ ¼

h i WðλÞ 1−10−εðλÞ Ci Z

ð9Þ

Therefore, the photolysis rate can be calculated using the equations ð10Þ

And X dC kp ðλÞCi − i¼ dt

ð11Þ

where kp(λ) is the photolysis rate (10 s−1) and ∅ is the NDMA quantum yield (0.41 ± 0.02 according to Plumlee and Reinhard, 2007). The molar extinction coefficient (ε(λ)) of NDMA was determined by Plumlee and Reinhard (2007). Absorbance (α(λ)) by the source water, determined in our laboratory, is shown in Fig. S.3. Absorbance by the biofilm tank effluent was almost the same as absorbance by the source water. Plumlee and Reinhard (2007) determined a photolysis rate (kp) for NDMA in Milli-Q water of 0.04 min−1 in 765 W/m2 natural sunlight. Solar irradiation data for DWTP7 were obtained from a NASA database (NASA Surface Meteorology and Solar Energy Website, n.d.) and the Code for the thermal design of civil buildings (Ministry of UrbanRural Development and Environmental Protection of the PRC, 1986. Midday solar irradiation (I0) at DWTP7 was 967 W/m2 in August and 387 W/m2 in December. Irradiation data for specific wavelengths were calculated from the solar spectrum distribution (Environmental testing, 2013). The sunlight wavelength from 290 to 450 nm was calculated in this model, for no absorption at higher wavelength according to Fig. S.3. 3.4.2. Simulating NDMA removal by photolysis The photolysis model was used to simulate the actual removal of NDMA in the biofilm and sedimentation tanks. The NDMA concentration in the sedimentation tank varied with water depth. An infinitesimal method was used to divide the sunlight wavelength into 10 nm intervals and the water depth into 0.1 m intervals (z). Each wavelength (λ1) had a corresponding initial solar irradiation value (I01), water absorbance value (α1), and NDMA molar extinction coefficient (ε1).Solar irradiation at water depth Zn can be calculated using Eq. (12). In1 ¼ I01 ∙10−αD ðλÞ Z n

h i kp ðλÞ∝∅In1 ðλÞεðλÞ 1−10−αD ðλÞ z zαðλÞ

ZCi

kp ðλÞ ¼ ka ðλÞ∅

obtained using Eq. (13).

ð12Þ

According to Eqs. (8) and (12), the photolysis rate for depth Zn can be

ð13Þ

Eq. (14) can be used to simulate the results of an experiment by Plumlee and Reinhard (2007), in which I0 was 765 W/m2, Z was 10 cm, and Ci was 100 μg/L. h i I0 1−10−εðλÞ Ci Z kp ðλÞ∝∅ ZCi

ð14Þ

Eqs. (13) and (14) allowed the final photolysis rates and NDMA concentrations at different depths shown in Fig. 8 to be calculated. As shown in Fig. 8, direct photolysis in the sedimentation tank in August and December removed 16 and 9.4%, respectively, of the NDMA present. The aerated biofilm tank can be regarded as continually stirred, which makes the NDMA concentrations in the tank homogenous. The photolysis rates in August and December, calculated using Eq. (15), were 0.71 and 0.28 h−1, respectively. The total amounts of NDMA removed in the aerated biofilm tank in August and December were 51% and 25%, respectively. Z kp;V ðλÞ ¼

0

Z

kp ðλÞdZ

ð15Þ

3.5. Calculation of NDMA removal by the treatment train The actual and calculated removals of NDMA in DWTP7 are shown in Table 2. It indicates that the biofilm pretreatment tank played the most important role in removing NDMA (48% in August and 22% in December), followed by the horizontal sedimentation tank. Sand filtration did not contribute to NDMA removal. Modeling results revealed that in the biofilm pretreatment tank, photolysis removed 51% of the total amount of NDMA in summer and 25% in winter. The biotreatment removed 0.2–12% of the total amount of NDMA in summer and 0.1–6.1% in winter. Aeration in the biofilm tank removed only 0.8% of the total amount of NDMA. These mechanisms fully explained the observed NDMA removal. In the horizontal sedimentation tank, modeling results showed that photolysis removed 16% of the total amount of NDMA in summer and 9.4% in winter. Volatilization in the sedimentation tank contributed insignificantly to NDMA removal. These mechanisms fully explained the NDMA removal.

Fig. 8. N nitrosodimethylamine (NDMA) (a) photolysis rates and (b) concentrations in the sedimentation tank.

Y. Qiu et al. / Science of the Total Environment 697 (2019) 133993 Table 2 Actual and calculated N nitrosodimethylamine removal efficienciesa in DWTP7.

Biofilm tank Sedimentation tank Sand filter a

Sampling time

Actual Calculated removal efficiency removal By By By aeration photolysis biodegradation or volatilization

Summer Winter Summer Winter Summer Winter

48% 22% 19% 9.2% – −1.7%

0.8% 0.02% –

51% 25% 16% 9.4% –

0.2–12% 0.1–6.1% – –

Calculated based on the influent and effluent NDMA concentration of each tank.

4. Conclusions and suggestion NDMA present in some Chinese source waters could be attenuated in the DWTPs. Here, a field study of a conventional DWTP with biopretreatment, as well as laboratory experiments, was conducted to quantify the contributions of volatilization, photolysis, and biodegradation. The following conclusions were obtained: (1) According to sampling results, a large removal of NDMA (48% in August and 22% in December) was observed in the biofilm pretreatment tank, followed by that in the horizontal sedimentation tank (19% in August and 9.2% in December). Negligible NDMA removal was observed in the sand filter. (2) According to modeling calculated results, photolysis played the most important role in removing NDMA in this DWTP, whereas volatilization presented insignificant removal. The contribution of biotreatment needs further investigation. The large NDMA removal in the bio-pretreatment tank was mainly due to photolysis and less due to biodegradation. (3) Photolysis, volatilization, and biodegradation could fully explain the NDMA removal in the DWTP studied. (4) We established a photolysis model based on theory and literature data. In the future, we intend to test the model under controlled laboratory conditions. Only preliminary work on NDMA biodegradation was performed in this study. The relationships between the biomass, microbial community structure, and NDMA biodegradation rate need to be further investigated.

Acknowledgements This research was supported by the National Natural Science Foundation of China (Grant No. 21777079), National Water Major Project (Grant No. 2018ZX07111-006), Tsinghua University Initiative Scientific Research Program (Grant No. 20173080012), National Geographic Air & Water Conservation Fund (Grant #GEFC-07-16), the Committee of Science and Technology Innovation of Shenzhen China (JCYJ20170817161942307) and the open project of State Key Joint Laboratory of Environmental Simulation and Pollution Control (16Y01ESPCT, 19Y02ESPCT). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2019.133993. References Afzal, A., Kang, J., Choi, B.M., Lim, H.J., 2016. Degradation and fate of N nitrosamines in water by UV photolysis. Int. J. Greenhouse Gas Control 52, 44–51. General Administration of Quality Supervision, Inspection and Quarantine of the PRC, Environmental testing - Part 2: Test methods-Test Sa: Simulated solar radiation at

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