Impact of upgrading wastewater treatment plant on

Science of the Total Environment 603–604 (2017) 140–147

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

Impact of upgrading wastewater treatment plant on the removal of typical methyl, oxygenated, chlorinated and parent polycyclic aromatic hydrocarbons Meng Qiao, Wei Cao, Bochuan Liu, Yaohui Bai, Weixiao Qi, Xu Zhao ⁎, Jiuhui Qu Key Laboratory of Drinking Water Science and Technology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 100085, China

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

• Chlorinated PAHs exist in WWTPs in Beijing with low concentration. • SPAHs should not be ignored due to the similar total concentration to PAHs. • PAHs and SPAHs mainly removed by secondary treatment but not tertiary treatment. • Dissolved organic matter might prevent the removal of PAHs and SPAHs in WWTPs. • Upgrading of WWTPs may not efficiently decrease the discharge of PAHs and SPAHs.

a r t i c l e

i n f o

Article history: Received 8 March 2017 Received in revised form 12 June 2017 Accepted 12 June 2017 Available online xxxx Editor: Adrian Covaci Keywords: PAHs SPAHs Secondary treatment Tertiary treatment Upgrading

⁎ Corresponding author. E-mail address: [email protected] (X. Zhao).

http://dx.doi.org/10.1016/j.scitotenv.2017.06.097 0048-9697/© 2017 Published by Elsevier B.V.

a b s t r a c t Wastewater treatment plant (WWTP) secondary effluent is a main source for polycyclic aromatic hydrocarbons (PAHs) and their derivatives (SPAHs) to wastewater receiving rivers in Beijing. The treatment technologies are being upgraded in the WWTPs as the tertiary treatment. To assess the improvement of the removal efficiencies of PAHs and SPAHs after the treatment upgrading, we investigated 16 PAHs and 4 types of SPAHs in the secondary and tertiary treatment process in 5 major WWTPs. Most of the parent PAHs, methyl PAHs, oxygenated PAHs and chlorinated PAHs were detected in the influent, secondary and tertiary effluent. The concentrations of ΣSPAHs (61 ng/L–529 ng/L) were similar to ΣPAHs (89 ng/L–474 ng/L), indicating that SPAHs should not be ignored when studying the PAH contamination. ΣPAHs and ΣSPAHs were largely removed by the secondary treatment (45%–82%) and less by the tertiary treatment (0%–24%). The removal efficiencies were lower in the secondary and tertiary treatment in WWTPs than in the lab-scale experiment conducted previously, probably a result of the association of PAHs and SPAHs with dissolved organic matters (DOMs) in wastewater. DOMs might be a limiting factor for the removal of PAHs and SPAHs in WWTPs. The estimated yearly loadings of the total PAHs and SPAHs decreased only 21% in the tertiary effluent compared with the secondary effluent in WWTP1 and 9% in WWTP3. Therefore, the upgrading of WWTPs did not efficiently improve the removal of PAHs and SPAHs. DOMs should be further considered for improving the removal of PAHs, SPAHs and similar contaminants in WWTPs. © 2017 Published by Elsevier B.V.

M. Qiao et al. / Science of the Total Environment 603–604 (2017) 140–147

1. Introduction Wastewater treatment plant (WWTP) effluent is a main source for polycyclic aromatic hydrocarbons (PAHs) and some of their derivatives (SPAHs) in wastewater receiving rivers in Beijing (Qi et al., 2013; Qiao et al., 2014b). Until now, the secondary treatment (biological treatment) effluent directly discharges to the receiving rivers. During the biological treatment process in WWTPs, the occurrence and removal of PAHs have been well studied previously (Bergqvist et al., 2006; Fatone et al., 2011; Manoli and Samara, 1999; Mezzanotte et al., 2016; Ozaki et al., 2015). However, to the best of our knowledge, SPAHs, such as methyl PAHs (MPAHs), oxygenated PAHs (OPAHs), polychlorinated naphthalenes (PCNs) and chlorinated PAHs other than PCNs (ClPAHs) have been seldom reported in WWTPs. The toxicities of some SPAHs, especially some ClPAHs, are higher than their corresponding PAHs (Bhatia et al., 1987; Iino et al., 1999; Ohura, 2007). ClPAHs and PCNs, emitted from traffic and incineration facilities, are widely spread in the atmosphere environment (Ohura, 2007). Similar to OPAHs, ClPAHs could be formed by PAHs during chemical reactions, such as PAHs incineration at 400–600 °C in the existence of O2 or Cl (Iino et al., 1999; Yoshino and Urano, 1997). ClPAHs have also been detected in tap water, formed during water chlorination (Johnsen et al., 1989; Shiraishi et al., 1985). PCNs have been detected in activated sludge from WWTPs (Stevens et al., 2003; Zhang et al., 2014). So, we suppose that ClPAHs and PCNs exist in the dissolved phase of the wastewater from WWTPs. MPAHs and OPAHs were detected during the secondary treatment in WWTPs in one study from our group (Qiao et al., 2014a). In addition, OPAHs were proved to be formed from PAHs during activated sludge treatment process (Qiao et al., 2016b). Therefore, besides PAHs, the occurrence and removal of SPAHs should be taken into consideration in wastewater treatment process. The treatment technologies are being upgraded in some of the WWTPs in Beijing (tertiary treatment). The tertiary treatment effluent will discharge into the receiving rivers instead of the secondary effluent in the future years. Therefore, it is necessary to investigate the removal of the toxic organic pollutants during the tertiary treatment process to forecast the improvement of water quality by the upgraded treatment. The removal of pharmaceutical and personal care products has been well studied during the tertiary treatment process, as well as in the WWTPs in Beijing (Hollender et al., 2009; Li et al., 2013; Sui et al., 2010). Ozonation is a commonly used technique in the tertiary treatment process, especially in the WWTPs in Beijing. PAHs ozonation has been studies in lab-scale experiment. PAHs were found to be easily eliminated by ozone (Trapido et al., 1995). However, in WWTPs, the removal of PAHs and SPAHs has rarely been investigated in the tertiary treatment process including the ozonation process. The objective of this study was to: 1) identify the existence of typical PCNs and ClPAHs in WWTPs; 2) investigate the removal efficiency and deduce the removal mechanism of PAHs and typical SPAHs both in the secondary and tertiary treatment processes; 3) estimate the discharge reduction of PAHs and SPAHs to the wastewater receiving rivers after the treatment technology upgraded. The result will provide bases for forecasting the improvement for the water quality of the WWTP effluent receiving river on the aspect of PAHs and typical SPAHs after the techniques upgraded in the WWTPs.

2. Materials and methods 2.1. Wastewater treatment process The removal of PAHs and typical SPAHs were investigated in 5 major WWTPs in Beijing by the secondary and tertiary treatment process. All the secondary treatment processes are biological treatment in the 5 WWTPs. The tertiary treatment processes are chemical treatment combined with biological or physical treatment in 3 WWTPs. The 5 WWTPs

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treat 70% of the domestic wastewater in Beijing. The details of the WWTPs are listed in Table 1.

2.2. Sample collection The influent, secondary effluent and tertiary effluent of the WWTPs were collected in April and November 2015, representing the nonheating season and the heating season respectively. The concentrations of PAHs and SPAHs were relatively stable on different days during one week based on our previous study (Qiao et al., 2014a). Additionally, during the sampling period in this study, there was no large precipitation which might influence the inflow velocity and quality. Therefore, the samples were collected once in each season. The sampling sites are shown in Table 1. Some samples could not be collected, including the secondary influent in WWTP3 in April, the tertiary effluent in WWTP4 and the secondary influent in WWTP5 in November. 4 L water was collected for each grab sample. The sampling was limited by the lack of auto samplers and also the sampling time limitation. Grab samples could be still deemed acceptable, because the concentrations of PAHs and SPAHs were relatively stable at 3 sampling times during one day period based on our previous study (Qiao et al., 2014a). Furthermore, as indicated above, during this sampling period, the inflow velocity and water quality were relatively stable.

2.3. Chemicals and materials Seventeen individual SPAH standards included 4 ClPAH standards, 9-chlorophenanthrene (9-ClPhe, 50 μg/mL in isoocatane), 2-chloroanthrancene (2-ClAnt, 50 μg/mL in isoocatane), 1chloroanthraquione (1-ClAQ, solid, 98%), 9,10-dichloroanthracene (9,10-DClAnt, solid, 96%) purchased from Chiron (Norway); 5 PCN standards, 1-chloronapthalene (1-CN, solid, 95.5%), 2-chloronapthalene (2-CN, 100 μg/mL in methanol), 1,4-dichloronaphthalene (DCN, solid, 99.4%), 1,2,3,4-tetrachloronaphthalene (TeCN, solid, 100%) and octachloronaphthalene (OCN, 100 μg/mL in methanol); 4 MPAH standards, 2-methylnaphthalene (2-MN, solid, 100%), 1-methylfluoranthene (1-MF, 10 μg/mL), 2,6-dimethylnaphthalene (2,6-DMN, solid, 100%), 3,6-dimethylphenanthrene (3,6-DMP, solid, 100%); 4 OPAH standards, 9-fluorenone (9-FL, solid, 100%), anthraquinone (AQ, 100 μg/mL), 2methylanthraquinone (2-MAQ, solid, 99.0%), benz[a]anthracene-7,12dione (BA-7,12-D, 50 μg/mL) purchased from AccuStandard, Inc. (USA). The sixteen USEPA priority PAHs, including naphthalene (Nap), acenaphthylene (Acy), acenaphthene (Ace), fluorene (Fluo), phenanthrene (Phe), anthracene (Ant), fluoranthene (Flua), pyrene (Pyr), benz[a]anthracene (BaA), chrysene (Chry), benzo[b]fluoranthene (BbF), benzo[k]fluoranthene (BkF), benzo[a]pyrene (BaP), indeno[1,2,3cd]pyrene (IcdP), dibenz[a,h]anthracene (DBA), and benzo[g,h,i]perylene (BghiP), in a mixture (200 μg/mL) were purchased from AccuStandard, Inc. (USA). The physical and chemical properties are listed in Table S1. Surrogate standards including 4 deuterated PAHs (d-PAHs) acenaphthene-d10 (d-Ace), phenanthrene-d10 (d-Phe), chrysene-d12 (d-Chry) and perylene-d12 (d-Pery) in a mixture (2000 μg/mL) were also purchased from AccuStandard, Inc. Internal standards including 2-fluorobiphenyl (2-FB, in solid N 96%) and decachlorobiphenyl (PCB209) were obtained from Aldrich Chemical Co., Inc. (Gillingham, Dorset, UK). Hexane (HEX, Fisher Scientific, USA), dichloromethane (DCM, J. T. Baker, USA) and methanol (MeOH, Fisher Scientific, USA) were HPLC grade solvents. Silica gel (0.06–0.2 mm) and alumina (100–200 mesh) for chromatography were purchased from Acros Organics, Inc. (USA). Before use, silica gel and alumina were baked at 180 °C and 250 °C respectively for 12 h, and then deactivated with 3% ultrapure water, and kept in HEX until use. Analytical grade anhydrous sodium sulfate (Tianjin, China) was baked at 450 °C for 5 h before use.

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Table 1 Information of the 5 WWTPs and sampling sites. WWTPs

WWTP1

WWTP2

WWTP3

WWTP4

WWTP5

Service population Secondary treatment process

814,000 Anaeroxic-anoxic-oxic/ anoxic-anaeroxic-oxic 550,000

2,400,000 Traditional activated sludge

500,000 Oxidation ditch activated sludge

400,000 MBR

2,415,000 Anaeroxic-anoxic-oxic

1,000,000

200,000

100,000

600,000

Membrane + ozone

–

MBR + ozone

–

550,000

–

Two stage biological filter + cloth filter + ozone 200,000

100,000

–

11 13 Apr.-influent Apr.-secondary effluent Apr.-tertiary effluent Nov.-influent Nov.-secondary effluent Nov.-tertiary effluent

11 20 Apr.-influent Apr.-secondary effluent – Nov.-influent Nov.-secondary effluent –

15 13 ××× Apr.-secondary effluent Apr.-tertiary effluent Nov.-influent Nov.-secondary effluent Nov.-tertiary effluent

17 20 Apr.-influent Apr.-secondary effluent Apr.-tertiary effluent Nov.-influent Nov.-secondary effluent ×××

11 15 Apr.-influent Apr.-secondary effluent – ××× Nov.-secondary effluent –

Secondary treatment capacity (m3 day−1) Tertiary treatment process Tertiary treatment capacity (m3 day−1) HRT (h) SRT (d) Sampling site

2.4. Analytical procedure The details of the analytical procedure were reported in previous study (Qiao et al., 2017; Qiao et al., 2013). A brief description is listed as follows. After the water samples transported to the laboratory, they were filtered with glass microfiber filters. The dissolved phase samples were extracted with solid phase extraction (SPE) method using C18 columns. The C18 columns were conditioned with 5 mL DCM, 5 mL MeOH and 5 mL purified water successively before extraction. All the samples were extracted within 2 days. After the columns were dried under vacuum for 0.5 h, the columns eluted with 10 mL DCM and 5 mL HEX successively. Then the crude extracts were solvent exchanged to HEX and concentrated to 1 mL. The silica gel and alumina were packed in columns (12 cm and 6 cm) for purification. The 1-mL crude extract was added to the purification columns. Then, Fraction 1 to Fraction 4 (F1-10 mL HEX, F2–75 mL 3:7 DCM:HEX, F3-60 mL 1:1 DCM:HEX, F460 mL 7:3 DCM:HEX) were used for elution, successively. F2–F4 was collected. The eluted fractions were concentrated to 0.5 mL and the internal standards were added for instrument analysis. PAHs and SPAHs in the particulate phase were mainly removed along with the particles. The contents of particles were relatively low in the secondary and tertiary effluent. The concentrations of PAHs and SPAHs in the particulate phase of the secondary effluent were demonstrated to be lower than those in the dissolved phase (Qiao et al., 2016a), which might not significantly influence the results and conclusion. Therefore, we only considered PAHs and SPAHs in the dissolved phase but not in the particulate phase samples in this study. An Agilent 7890A gas chromatography equipped with a 5975C mass detector (GC–MS) with electron impact (EI) source in selected ion monitoring (SIM) mode were used for the detection. A DB-17MS fused silica capillary column (30 m × 0.25 mm × 0.25 μm) was used for the compounds separation. Samples (1 μL) were injected in splitless mode. The carrier gas was helium. The injector and detector temperatures were 280 °C and 290 °C, respectively. The temperature program procedure was: 60 °C (1 min), to 110 °C at a rate of 20 °C/min, and to 290 °C at a rate of 3 °C/min (20 min).

2.5. Quality assurance and quality control 100 ng for each d-PAH was added to each water sample after filtration and before extraction to confirm the recoveries for each sample. The recoveries of the d-Ace, d-Phe, d-Chry and d-Pery were 53% ± 13%, 66% ± 4%, 89% ± 15% and 83% ± 16%, respectively. 250 ng for each target compound was added to the real matrix sample to confirm the recoveries of all the target compounds. The secondary effluent from WWTP1 in April was selected as the real matrix sample. Three 4-L

samples were used for recovery calculation in triplicate. The recoveries were calculated as follows: Re ð%Þ ¼ ðCadd −Cori Þ 100=250 where Cadd was the content of the target compounds in the real samples with adding the target compounds; Cori was the content of the target compounds in the real samples without adding the target compounds. The method detection limits (MDLs) and the method quantification limits (MQLs) were calculated based on the 3 and 10 time ratio of the signal to noise, respectively. The laboratory background concentrations were measured using 4-L ultrapure water through the whole extraction and purification procedure in triplicate. The recoveries, the MDLs and MQLs, and the background concentrations of all the target compounds are listed in Table S2. 2.6. Calculation of removal efficiency The removal efficiencies during the secondary and tertiary treatment process were calculated as follows: Rsec ð%Þ ¼ ðCin −C sec−eff Þ 100=Cin Rter ð%Þ ¼ ðC sec−eff −Cter−eff Þ 100=C sec−eff Rter−total ð%Þ ¼ ðC sec−eff −Cter−eff Þ 100=Cin where Cin, Csec-eff and Cter-eff were the concentrations of the target compounds in the influent, secondary and tertiary effluent, respectively. 3. Results and discussion 3.1. Occurrence of PAHs and SPAHs in the influent and effluent of the WWTPs The detection rates of most PAHs were 100% in the influent, secondary and tertiary effluent both in April and November, except IncdP and DBA (Table 2). No IncdP was detected in the secondary and tertiary effluent in April. No DBA was detected in all the WWTPs. Regarding SPAHs, the detection rates of MPAHs and OPAHs were 100%, in accordance with our previous studies (Qiao et al., 2014a). The detection rates of most ClPAHs were 100%, except 9,10-DClAnt (0%). Regarding PCNs, 2-CN and DCN were 100% detected; 1-CN was 100% detected in the influent in April; TeCN and OCN were not detected. Regarding individual compounds, the concentrations of Nap, Phe, 2-MN, 9-FL, AQ and 2-MAQ were higher than the other PAHs and SPAHs (Fig. S1). The existences of the low-ring-number PAHs, MPAHs, OPAHs, PCNs and ClPAHs were more prevalent than the high-ring-number compounds,


143

Table 2 Detection rates of the 33 PAH and SPAH compounds in the influent, secondary and tertiary effluent in WWTPs. %

PAHs

MPAHs

OPAHs

ClPAHs

PCNs

April 2015

Nap Acy Ace Fluo Phe Ant Flua Pyr BaA Chry BbF BkF BaP IncdP DBA BghiP 2-MN 2,6-DMN 3,6-DMP 1-MF 9-FL AQ 2-MAQ BA-7,12-D 9-ClPhe 2-ClAnt 9,10-DClAnt 1-ClAQ 2-CN 1-CN DCN TeCN OCN

November 2015

Influent

Secondary effluent

Tertiary effluent

Influent

Secondary effluent

Tertiary effluent

100 100 100 100 100 100 100 100 100 100 100 100 100 100 0 100 100 100 100 100 100 100 100 100 100 100 0 100 100 100 100 0 0

100 100 100 100 100 100 100 100 100 100 100 100 100 0 0 100 100 100 100 100 100 100 100 100 100 100 0 100 100 0 100 0 0

100 100 100 100 100 100 100 100 100 100 100 100 100 0 0 100 100 100 100 100 100 100 100 100 100 100 0 100 100 0 100 0 0

100 100 100 100 100 100 100 100 100 100 100 100 100 100 0 100 100 100 100 100 100 100 100 100 100 100 0 100 100 0 100 0 0

100 100 100 100 100 100 100 100 100 100 100 100 100 100 0 100 100 100 100 100 100 100 100 100 100 100 0 100 100 0 100 0 0

100 100 100 100 100 100 100 100 100 100 100 100 100 100 0 100 100 100 100 100 100 100 100 100 100 100 0 100 100 0 100 0 0

attributed to the lower log Kow values (Table S1) of the low-ringnumber compounds which were prone to exist in the dissolved phase. The existences of PCNs and ClPAHs were less prevalent than PAHs, similar to those in the atmosphere environment (Ishaq et al., 2003; Ohura, 2007), possibly because the discharges of ClPAHs and PCNs were less than PAHs to the environment. From another aspect, ClPAHs might be easier transformed than PAHs, resulting to a less prevalent of ClPAHs (Ohura et al., 2004). Generally, the concentrations of ΣPAHs (16 PAHs, 89 ng/L–474 ng/L) were higher than ΣOPAHs (4 OPAHs, 36 ng/L–264 ng/L), than ΣMPAHs (4 MPAHs, 7 ng/L–147 ng/L), than ΣClPAHs (4 ClPAHs, 17 ng/L–44 ng/L), than ΣPCNs (5 PCNs, 4 ng/L–25 ng/L) (Table 3). Though there still were other SPAH compounds that we did not consider, the concentrations of ΣSPAHs (61 ng/L–529 ng/L, ΣOPAHs + ΣMPAHs + ΣClPAHs + ΣPCNs) were similar to the concentrations of ΣPAHs (Fig. 1). Thus, SPAHs should not be ignored when studying the PAHs contamination. The concentrations of PAHs and SPAHs were higher in April (non-heating season) than in November (heating season). The result was different with previous studies that the concentrations of PAHs and SPAHs in the nonheating season were lower than in the heating season (Qiao et al., 2014a; Qiao et al., 2014b). Considering the emission sources of PAHs

and SPAHs, the higher concentrations of PAHs in winter were attributed to the coal combustion for heating (Wang et al., 2011b). In recent years, the natural gas gradually replaced coal combustion for heating as clean energy in winter in Beijing (Xinhuanet, 2015). Thus, the reduction of the coal used for heating probably reduced the PAHs concentrations in winter. Therefore, the seasonal variation of PAHs might not be controlled by the coal combustions for heating. Besides the heating in winter, the coal used for electric power generation was also reduced recently (Online, 2017). The 2016 statistical yearbook of Beijing City showed that from 2010 to 2015, the consumption of coal decreased, while the consumption of natural gas increased obviously, especially in 2015 (Fig. S2) (Beijing-Statistics-Bureau, 2016). 2015 is a special year for energy transition from coal to clear energy. Half of the major coal-fired power plants in Beijing had been switched off in March 2015. The higher concentrations in April were probably resulted from the memory emissions of PAHs after the combustion switched off (Zimmermann et al., 2001). In November, not only the coal used for heating, but also used for power generation were largely reduced compared with the earlier months in 2015, which probably resulted in the lower PAH concentrations in winter than spring in this special year. However, studies should be continued in the following years to confirm these phenomena.

Table 3 Concentrations of ΣPAH, ΣMPAHs, ΣOPAH, ΣClPAHs and ΣPCNs in the influent, secondary and tertiary effluent of WWTPs. ng/l

ΣPAHs ΣMPAHs ΣOPAHs ΣClPAHs ΣPCNs

April 2015

November 2015

Influent

Secondary effluent

Tertiary effluent

Influent

Secondary effluent

Tertiary effluent

474 ± 51 147 ± 56 264 ± 35 33 ± 4 25 ± 2

226 ± 26 32 ± 5 157 ± 9 18 ± 0 11 ± 2

161 ± 25 22 ± 13 132 ± 11 17 ± 0 8±1

333 ± 41 65 ± 33 199 ± 67 44 ± 3 15 ± 5

89 ± 14 7±4 38 ± 4 21 ± 2 4±1

98 ± 21 12 ± 4 36 ± 5 23 ± 3 6±1

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loadings in the influent of WWTP1 (200–350 μg capita−1 day−1) were higher than other WWTPs (80–181 μg capita−1 day−1), though the concentrations in WWTP1 were not the highest, indicating that there might be other sources for the influent other than domestic wastewater, such as industrial wastewater. The loadings in the secondary and tertiary effluent of the WWTPs ranged from 15 to 144 μg capita−1 day−1 and from 27 to 102 μg capita−1 day−1. In a WWTP in Japan from 2005 to 2006, the daily loadings of PAHs in the influent and effluent was 142 ± 53 and 28 ± 11 μg capita−1 day−1, respectively. It should be noted that the daily loadings in Japan was in the total phase, while in this study was in the dissolved phase (Ozaki et al., 2015). So if only compared the dissolved phase, the daily loadings in this study should be higher than those in Japan. In the WWTPs in Canada during 1993, the daily loadings of 21 PAHs in the influent and effluent were 1786 and 500 μg capita−1 day−1, possibly attributed to direct discharge of the industrial wastewater to the sewer system (Pham and Proulx, 1997). The daily loadings of PAHs in the WWTPs in Canada in the end of last century were much higher than those in this study. The result might reflect that during the last 2 decades, coal combustions for industry and domestic use have been gradually transformed to clean energy in most of the developed and developing countries. 3.2. Removal of PAHs and SPAHs in the secondary and tertiary treatment

Fig. 1. Daily loading per capita and concentration of ΣPAHs and ΣSPAHs in different WWTPs (cycle stands for PAHs; triangle stands for SPAHs).

The concentrations of ΣPAHs in the influent, secondary effluent and tertiary effluent were 333 ± 41 ng/L–474 ± 51 ng/L, 89 ± 14 ng/L–226 ± 26 ng/L and 98 ± 21 ng/L–161 ± 25 ng/L in dissolved phase, respectively (Table 3). The influent concentrations of ΣPAHs in WWTPs of this study were much lower than those in 2 Italy WWTPs during 2011 and 2013 (4770 ± 3740 ng/L in Alto Seveso and 2110 ± 1120 ng/L in Nosedo in dissolved phase) (Mezzanotte et al., 2016). The concentrations in the total phase but not the dissolved phase of the secondary and tertiary effluent were reported in the Italy WWTPs. The contributions of the particulate phase were much less than the dissolved phase to the total phase (dissolved + particulate phase) according to the influent. So the total phase could be compared with the dissolved phase in this study. In the total phase in the Italy WWTPs, the concentrations of ΣPAHs were 190 ± 160 ng/L–260 ± 160 ng/L in the secondary effluent, and 150 ± 130 ng/L–150 ± 70 ng/L in the tertiary effluent, which were similar to those in this study. In Japan, the concentrations of ΣPAHs in the influent and secondary effluent were 64 ± 48 ng/L and 32 ± 35 ng/L in the dissolved phase, which were much lower than those in this study. The concentrations of ΣPAHs, ΣMPAHs and ΣOPAHs in 2015 were compared with those in the same WWTPs in previous years (Fig. S3) (Qiao et al., 2016a; Qiao et al., 2014a; Qiao et al., 2014b). The concentrations of ΣPAHs and ΣMPAHs decreased both in the influent and effluent from 2010 to 2015, demonstrating again a less pollution situation of PAHs and MPAHs in recent years. The concentrations of ΣOPAHs did not vary much during 2010 and 2015, since that OPAHs in the environment could be formed by the biological and chemical transformation from PAHs, besides the burning of coal (Lundstedt et al., 2006). The daily mass loadings per capita of ΣPAHs and ΣSPAHs were calculated by: (Concentration: ng/L) × (Daily treatment capacity: m3 day−1) / (Population: capita−1), illustrated in Fig. 1. The mass loadings of PAHs per capita were usually calculated in the influent and effluent of WWTPs (Barrick, 1982; Ozaki et al., 2015; Pham and Proulx, 1997). The loadings in the influent reflected the personal discharge of the pollutants in the service area; the loadings in the effluent were used for estimating the contribution to the receiving environment. It could be noted that the

The removal efficiencies of ΣPAHs and ΣSPAHs during the secondary treatment process in November (67%–82% and 70%–82%) were significantly higher than in April (45%–59% and 48%–59%) (Paired t-test, P b 0.05). On the contrary, the removal efficiencies of ΣPAHs and ΣSPAHs during the tertiary treatment process in April (10%–22% and 4%–24%) were significantly higher than in November (0% and 0%) (Paired t-test, P b 0.05). Temperature might be a significant factor influencing the removal efficiencies in different seasons. The adsorption of PAHs and SPAHs would be promoted at the lower temperature in the range of 10 °C–30 °C (Qiao et al., 2016b), due to a higher linear sorption coefficient (Kd) of the compounds at the lower temperature (Ren et al., 2007; Wang et al., 2011a). The biotransformation would be promoted at the higher temperature in the range of 10 °C–30 °C (Qiao et al., 2016b), resulted from more plentiful microbial species at the higher temperature (Wu, 2005). Therefore, the higher removal efficiencies in November than in April demonstrated that adsorption was a main removal mechanism. The removal efficiencies of different ring number PAHs and SPAHs were illustrated in Fig. 2. It could also be noticed that the removal efficiencies generally decreased with the increased ring number for PAHs, MPAHs, OPAHs and ClPAHs including PCNs in the secondary treatment process, both in April and November. The biodegradation efficiency decreased and the adsorption efficiency increased with the increasing ring number (molecular weight) of PAHs (Qiao et al., 2016b). The result indicated that the target compounds were also removed through biotransformation in the secondary treatment process. Consequently, the removal mechanism of PAHs and SPAHs in the secondary treatment process was both by adsorption and biotransformation, consistent with previous studies (Manoli and Samara, 1999; Qiao et al., 2014a). The tertiary treatment processes were combined physical (ultrafiltration) or biological technique (biofiltration, MBR) with chemical technique (ozonation). The ultrafiltration process should be inefficient for the removal of PAHs and SPAHs because the molecular weights of PAHs and SPAHs (b 300 Da) were much lower than the ultrafiltrable compounds (1000 Da) (Sui et al., 2010). Aerated biofiltration was efficient for the removal of two- and three-ring PAHs (80%–100%), but not for the four-ring compounds (fluoranthene and pyrene) (Richard and Dwyer, 2001). The removal efficiencies of PAHs by the membrane bioreactor varied much, from 19% to 92% in different WWTPs and the removal efficiencies decreased with the increasing ring number (Fatone et al., 2011). So these processes seemed not play the main role for the removal of PAHs and SPAHs because the removal


120 100

WWTP1-Sec WWTP1-Ter

145

Apr. 2015

80

40 20 0 2R -P A 3R H -P A 4R H -P A 5R H -P A 6R H -P 2R AH -M P 3R AH -M P 2R AH -O P 3R AH -O P 4R AH -O PA H PC 3R N -C lP AH

Contribution of removal efficiency (%)

60

120

Nov. 2015

100 80 60 40 20 0

2R -P A 3R H -P A 4R H -P A 5R H -P A 6R H -P 2R AH -M P 3R AH -M P 2R AH -O P 3R AH -O P 4R AH -O PA H PC 3R N -C lP AH

-20

Fig. 2. Removal efficiencies in the secondary and tertiary treatment process during different seasons (R-ring, Sec-secondary treatment process, Ter-tertiary treatment process).

Fig. 3. Contributions of the secondary and tertiary treatment to the total removal efficiencies.

efficiencies in this study did not significantly varied with the ring number (Paired t-test, P N 0.05). Ozonation is the common technique existed in all the 3 tertiary treatment processes. It was reported that the ozone degradation efficiencies of PAHs influenced by the ring numbers (Trapido et al., 1995), similar to the result in this study. In addition, the chemical reaction might be enhanced by a higher temperature, resulting to the higher removal efficiencies in April than in November. Thus, PAHs and SPAHs in the tertiary treatment process might be mainly removed through ozone degradation. Furthermore, taken WWTP1 as an example, the contributions of the secondary and tertiary treatment to the total treatment removal efficiency were compared (Fig. 3). The compounds were largely removed by the secondary treatment process and less by the tertiary treatment process in the both seasons. The result was similar to a WWTP in Italy (Mezzanotte et al., 2016), but different from the removal of pharmaceuticals in the same WWTP in Beijing. Most of the pharmaceutical investigated could be efficiently removed during the ozonation treatment rather than during the biological secondary treatment (Sui et al., 2010). Similar result was obtained in a WWTP in Switzerland (Hollender et al., 2009). During the secondary treatment process, the compounds firstly experience adsorption and then biodegradation. The adsorption ability of the compounds were correlated to the log Kow values. Compared with PAHs and SPAHs, the log Kow values of pharmaceuticals were lower (Sui et al., 2010), resulting to a lower adsorbance on the activated sludge. It was reported that sulfonamide antibiotics biodegradation ‘was inhibited in the first 12 h possibly due to competitive inhibition of xenobiotic oxidation by readily biodegradable substances’ (Yang et al., 2011), and would be totally removed in 14 days (Yang et al., 2012). The hydraulic retention times of the investigated WWTPs were 11–18 h. Thus, during the staying time in WWTPs, certain pharmaceuticals could only be partly removed. In comparison,

PAHs and SPAHs with higher log Kow values were easier adsorbed on the activated sludge than pharmaceuticals, leading to a higher removal efficiencies during the secondary treatment process. It was different in the tertiary treatment. Ozonation was efficient for PAHs degradation in lab-scale experiments, during which PAHs could be totally removed in a few minutes (Sakulthaew et al., 2015; Trapido et al., 1995); whereas in this study, the removal efficiencies PAHs and SPAHs were relatively low during the ozonation process (0%–24%). Similar result was found in the Italian WWTP mentioned above, where the removal efficiency of PAHs during ozonation treatment was 0.89% (Mezzanotte et al., 2016). The low removal efficiencies of PAHs in the ozonation process in real WWTPs might be influenced by the dissolved organic matters (DOMs) in the dissolved phase. It was reported that the removal of DOC and UV254 in the secondary effluent by ozonation was relatively low (15% and 36%) (Zheng et al., 2014). PAHs and SPAHs were hydrophobic organic compounds with high log Kow values, prone to associated with the effluent organic matter (EfOM) (Jin et al., 2016). Thus, the association of PAHs with EfOM might reduce the removal efficiencies of PAHs in ozonation treatment in real WWTPs. In comparison, the higher removal efficiencies of pharmaceuticals in ozonation treatment probably resulted from the lower log Kow values. The DOMs might not affect the ozonation of pharmaceuticals. Therefore, the contributions of the tertiary treatment for PAHs and SPAHs removal were much less than the pharmaceuticals to the total removal efficiency. Regarding the influence of DOMs to the ozonation treatment in real WWTPs, similar phenomenon was obtained in the biological treatment. In a lab-scale experiment, the added PAHs were found to be more efficiently removed than the inherent PAHs (Qiao et al., 2016b), probably because DOMs in the dissolved phase were associated with PAHs. The association time of DOMs with the inherent PAHs was much longer than with the added ones. So the inherent PAHs were prone to exist in

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dissolved phase with DOMs and resulting to a lower adsorption ability and bioavailability on the activated sludge. As a result, the removal efficiencies of PAHs and SPAHs in the lab-scale experiment were higher than in WWTPs due to the influence of DOMs in WWTPs. Therefore, DOMs were probably the limiting factor for the removal of PAHs and SPAHs both in the secondary and tertiary treatment process. However, this speculation has not been proved during the ozonation process, and should be further confirmed in lab-scale experiment. 3.3. Estimating the yearly discharge decrease of PAHs and SPAHs to the receiving rivers after the treatment upgrading In order to further predict the improvement of PAHs and SPAHs pollution in the receiving rivers after the treatment upgraded, the PAHs and SPAHs yearly loadings were estimated in the secondary and tertiary effluent. Yearly loading (g) = [(CApr.: ng/L) + (CNov.: ng/L)] × (Daily treatment: m3/d) × (365: d) × 10− 3 / 2, where CApr. and CNov. were the concentration of Σ(PAHs + SPAH) in April and November, respectively. The yearly loadings of the total PAHs and SPAHs in the secondary effluent of WWTP1 and WWTP3 were 62 g and 21 g, and in the tertiary effluent were a little lower (49 g and 19 g) (Fig. 4). As a result, 21% yearly loading of PAHs and SPAHs decreased in the tertiary effluent compared with the secondary effluent in WWTP1; and 9% decreased in WWTP3. Consequently, the upgrading of the wastewater treatment plant was seemed not to be quite efficiently improved the water quality of the effluent receiving rivers from the aspect of PAH and SPAH contamination. 4. Conclusion Most of the parent PAHs, MPAHs, OPAHs, ClPAHs and PCNs were detected in the dissolved phase of the influent, secondary and tertiary effluent in the 5 major WWTPs in Beijing. The concentrations of ΣClPAHs and ΣPCNs (17 ng/L–44 ng/L and 4 ng/L–25 ng/L) were relatively low compared with ΣPAHs. The total concentrations of ΣSPAHs (61 ng/L– 529 ng/L) were similar to ΣPAHs (89 ng/L–474 ng/L), indicating that SPAHs should not be ignored when studying the PAHs contamination. ΣPAHs and ΣSPAHs were largely removed by the secondary treatment process (45%–82%) and less by the tertiary treatment process (0%–24%) in the both seasons. The removal mechanism of PAHs and SPAHs in the secondary treatment process was both by adsorption and biotransformation, and in the tertiary treatment process might mainly through ozone degradation. The removal efficiencies of PAHs and SPAHs were lower in WWTPs than in the lab-scale experiment previous studied, possibly caused by the association of PAHs and SPAHs with DOMs in the wastewater. DOM was probably a limiting factor for

Fig. 4. Estimated yearly loadings of the total PAHs and SPAHs in the secondary and tertiary effluent.

the removal of PAHs and SPAHs both in the secondary and tertiary treatment process. The estimated yearly loadings of the total PAHs and SPAHs decreased 21% in the tertiary effluent compared to the secondary effluent in WWTP1 and 9% in WWTP3. Therefore, the upgrading of the wastewater treatment plant might not be efficiently improved the water quality of the effluent receiving rivers on the aspect of PAH and SPAH contamination. So, further studies should be considered on DOM to increase the removal efficiencies of PAHs, SPAHs and similar hydrophobic microorganic contaminants.

Acknowledgements This work was supported by National Natural Science Foundation of China (Grant No. 51508552 and 51420105012) and Key Program of the Chinese Academy of Sciences (Grant No. ZDRW-ZS-2016-5-6). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2017.06.097.

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