Tidal variability of polycyclic aromatic hydrocarbons and organophosphate esters in the coastal seawater of Dalian, China

Journal Pre-proof Tidal variability of polycyclic aromatic hydrocarbons and organophosphate esters in the coastal seawater of Dalian, China

Lijie Zhang, Yan Wang, Feng Tan, Ya Yang, Xiaowei Wu, Wei Wang, Dongmei Liu PII:

S0048-9697(19)34432-8

DOI:

https://doi.org/10.1016/j.scitotenv.2019.134441

Reference:

STOTEN 134441

To appear in:

Science of the Total Environment

Received date:

20 July 2019

Revised date:

11 September 2019

Accepted date:

12 September 2019

Please cite this article as: L. Zhang, Y. Wang, F. Tan, et al., Tidal variability of polycyclic aromatic hydrocarbons and organophosphate esters in the coastal seawater of Dalian, China, Science of the Total Environment (2019), https://doi.org/10.1016/ j.scitotenv.2019.134441

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© 2019 Published by Elsevier.

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Tidal variability of polycyclic aromatic hydrocarbons and organophosphate esters in the coastal seawater of Dalian, China Lijie Zhang a, Yan Wang

a, *

, Feng Tan a, Ya Yang a, Xiaowei Wu a, Wei Wang b,

Dongmei Liu b,* a

Key Laboratory of Industrial Ecology and Environmental Engineering (MOE),

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School of Environmental Science and Technology, Dalian University of Technology, Dalian 116024, China

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Technology, Harbin 150090, China

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State Key Laboratory of Urban Water Resources & Environment, Harbin Institute of

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b

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* Corresponding author. E-mail: [email protected]

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* Corresponding author. E-mail: [email protected]

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Abstract We investigated the tidal variability of polycyclic aromatic hydrocarbons (PAHs) and organophosphate esters (OPEs) in water dissolved phase from a coastal area of Dalian, China, as well as their air-water exchange trends. The concentrations of PAHs and OPEs in water were in the range of 50.5-74.7 ng/L and 21.6-61.5 ng/L,

(FLU)

for

PAHs,

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respectively. Phenanthrene (PHE) was the dominant congener followed by fluorene while

tris(2-chloroisopropyl)

phosphate

(TCIPP)

and

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tris(2-chloroethyl) phosphate (TCEP) dominated for OPEs. PAHs in coastal water

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showed a tidal variability, but not for OPEs, which may due to the influence of

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occasional wastewater discharges of OPEs. The source apportionments using

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principle component analysis and positive matrix factorization suggested that PAHs in

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the coastal water mainly came from oil spill from ships, coal combustion, and petroleum combustion, while OPEs were derived from diverse sources. The fugacity

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fractions (ff) suggested that ACY, ACE, FLU, PHE, TCEP, and TPHP volatilized from water into air, while TNBP, TCIPP, and TDCIPP deposited from air into water, and FLA, PYR, BaA, CHR, and EHDPP reached equilibrium. The ff values varied slightly with tidal circle, but the variations were not enough to alter the air-water exchange directions of those compounds. Although the influences of tide on the air-water exchange of PAHs and OPEs were limited, tide still played an important role on the transports and diffusions of those chemicals in the coastal water, which requires further studies.

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Keywords: Tidal variability; Coastal water; PAHs; OPEs; Air-water exchange

1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are a group of widespread contaminants, which have caused worldwide concern for decades. PAHs have been

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detected in various environmental matrices with anthropogenic sources dominating their environmental inputs, such as incomplete combustion of fossil fuel or biomass

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(Motelay-Massei et al., 2006). Due to the strong hydrophobicity together with the

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acute toxicity, carcinogenicity, and mutagenicity, PAHs can be accumulated to high

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levels in aquatic organisms and cause damage to the ecosystem (Li et al., 2009).

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Due to the phase out of polybrominated diphenyl ethers (PBDEs) (Kawagoshi et

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al., 2002), usage of organophosphate esters (OPEs) as flame retardants in industrial products has increased (Moller et al., 2012). The production of OPEs in China was

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~100 000 t in 2011, and increased by ~15% annually (Ou, 2011). The extensive use of OPEs has caused their high detection frequencies in coastal areas, including water, sediments, even biota samples (Cao et al., 2012; Moller et al., 2012; van der Veen and de Boer, 2012). Furthermore, halogenated OPEs, which dominate in the environment, are identified to be toxic and risky to human health (van der Veen and de Boer, 2012). Coastal area, a transitional and buffering zone between land and ocean (Zhang et al., 2004), is significantly influenced by human activities. Inputs of PAHs and OPEs to coastal water include either direct (e.g., sewage outflow, river discharge, traffic emission, and oil spill) or indirect (e.g., atmospheric deposition and air-water gas 3

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exchange) (Zhang et al., 2012; Ya et al., 2014). River flows and wastewater treatment plants have been proved to pose significant influences on concentrations of PAHs and OPEs in coastal areas (Lohmann et al., 2011; Wang et al., 2015). Meanwhile, since PAHs and OPEs are semivolatile compounds and occur in both gas and particle phases, they can be introduced into coastal water bodies by both dry and wet

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depositions (Jurado et al., 2005; Kim and Chae, 2016). Tidal current can lead to the exchange and mixing of fresh saline offshore water

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with inshore water, which is important for the transport and sedimentation of

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contaminant in oceans (Maskaoui et al., 2002). Hence, distribution and fate of organic

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contaminants in a tide-dominated area can be typically influenced by tidal mixing

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(Zhang et al., 2004; Liu et al., 2014). Previous studies suggested that tidal-driven

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mixing process could influence the concentrations (Liu et al., 2017) and dispersion of PAHs in a tide-dominated estuary (Maskaoui et al., 2002; Liu et al., 2018). However,

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knowledge regarding the influence of tide on concentrations and dispersions of PAHs and OPEs in coastal regions is still limited so far. Therefore, the major objectives of this study were to: (1) discover the concentrations and diurnal variations of PAHs and OPEs in the coastal water of Dalian, China; (2) identify the influence of tidal mixing and sources on the concentrations of PAHs and OPEs in this area; (3) estimate the variations of air-water exchange trends of PAHs and OPEs under the influences of tide and human activities.

2. Material and methods 4

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2.1 sampling Sampling was conducted in a coastal area of Dalian, China. Dalian is an important industrial port city located in the southern tip of Liaodong Peninsula between the Bohai Sea and the Yellow Sea. The sampling site is located on the Yellow Sea coast near an urban district. The coastal surface water samples (~4 L, ~0.5 m depth) were collected every 3 hours from 11:00 am September 21st, 2017 to 12:00 pm

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September 23th, 2017 according to the tidal process (17 samples in total). After

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sampling, the water samples were transported to the laboratory and analyzed

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immediately. The tidal height was recorded, and the corresponding water

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physicochemical properties, including dissolved organic carbon, dissolved nitrogen,

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water temperature, pH, dissolved oxygen, salinity, and conductivity, were measured

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simultaneously during sampling.

Two air samples were collected for 48 hours by a high-volume active air sampler

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located close to the water sampling site (<1 km) on September 18th and September 25th, 2017, respectively. Air sample was first passed through a Quartz fiber filters (Whatman, 20.3×25.4 cm, 450°C for 4 h), and then through a polyurethane foam (PUF) plug (6.5 cm diameter, 7.5 cm thickness, pre-cleaned by ethyl acetate and dichloromethane). A wireless weather station was set to record the meteorological parameters simultaneously. After sampling, the PUF samples were wrapped with aluminum foil, placed into polythene zip-bags and stored at -20°C until analysis. 2.2 Sample extraction and analysis Water sample was filtered by a glass fiber filter (GF/F, pre-baked at 450°C for 4 5

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h). Approximately 1 L of the filtered water was first spiked with surrogate standards, including naphthalene-d8 (NAP-d8), acenaphthylene-d8 (ACY-d8), phenathrene-d10 (PHE-d10), chrysene-d12 (CHR-d12), and perylene-d12 (PRY-d12) for PAHs, and tris (2-chloroethyl) phosphate-d12 (TCEP-d12), tris (2-chloroisopropyl) phosphate-d18 (TCIPP-d18), and tripropyl phosphate-d15 (TPHP-d15) for OPEs, and then liquid-liquid

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extracted using 20 mL dichloromethane (DCM) for 4 times. Anhydrous sodium sulfate (pre-baked at 450°C for 4 h) was used to remove water from the combined

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extracts. PUF plug was extracted by a Dionex ASE 350 system at 100°C for 5 min

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and 2 cycles using DCM: hexane (1:1 v/v). Extract was concentrated to ~0.5 mL after

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solvent-exchange to n-hexane, and purified by a silica gel column containing neutral

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silica gel (3% deactivated) and anhydrous sodium sulfate from bottom to top. The

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column was first eluted with 20 mL DCM/hexane (1:1, v/v) for PAHs, and then eluted with 20 mL ethyl acetate for OPEs. Hexamethyl benzene was added as the internal

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standard prior to instrumental analysis. 2.3 Instrumental analysis

Samples were determined for 15 PAHs, including acenaphthene (ACE), ACY, fluorene (FLU), PHE, anthracene (ANT), fluoranthene (FLA), pyrene (PYR), benz[a]anthracene (BaA), CHR, benzo[b]fluoranthene (BbF), benzo[k]fluoranthene (BkF), benzo[a]pyrene (BaP), indeno[1,2,3-cd]pyrene (IcdP), dibenz[ah]anthracene (DahA), benzo[ghi]perylene (BghiP), and 9 OPEs: including tri-n-buthyl phosphate (TNBP), TCEP, TCIPP (mix of 3 isomers), tris(2-chloro-1-(chloromethyl)ethyl) phosphate (TDCIPP), tris(2-butoxy)ethyl phosphate (TBOEP), TPHP, 2-ethylhexyl 6

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diphenyl

phosphate

(EHDPP),

tri(2-ethylhexyl)

phosphate

(TEHP),

and

tris(methylphenyl) phosphate (TMPP, mix of 4 isomers). The analysis was performed by an Agilent 6890GC-5975MSD using a DB-5MS capillary column (30 m×0.25 mm×0.25 μm) and an electron ionization impact (EI) source. For PAHs, the initial temperature was set at 80°C for 3 min, then increased to 290°C at 5°C/min, and held

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for 5 min. For OPEs, the temperature initialed from 70°C for 2 min, increased to 300°C at 15°C/min, and held for 10 min. Helium was used as the carrier gas at a flow

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2.4 Quality Assurance/Quality Control

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of 1 mL/min for PAHs and 1.5 mL/min for OPEs.

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A field blank and a procedure blank were analyzed simultaneously with every 8

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samples to assess potential contaminations. Method detection limit (MDL) was

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estimated as 3 folds of the standard deviation of blanks. The MDLs were 0.01-3.8 ng/L for PAHs and 0.15-4.5 ng/L for OPEs, respectively. The surrogate recoveries

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were 67.1 ± 14.9%, 86.2 ± 14.1%, 99.4 ± 12.9%, 103.8 ± 13.2%, and 96.5 ± 21.9% for NAP-d8, ACY-d8, PHE-d10, CHR-d12, and PRY-d12, and 79.3 ± 13.6%, 84.3 ± 15.4%, and 91.8 ± 10.5% for TCEP-d12, TCIPP-d18, and TPHP-d15, respectively. The results were corrected by both blanks and surrogate recoveries.

3. Results and discussion 3.1 PAHs and OPEs in seawater and air samples 3.1.1 PAHs in seawater The concentrations of BbF, BkF, BaP, IcdP, DahA, and BghiP in the water 7

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dissolved phase were mostly under MDL in this study, and were not discussed. The total concentrations of nine PAHs (Σ9PAHs) in the water phase ranged from 50.5 ng/L (23:00, 21st September) to 74.7 ng/L (06:00, 23rd September) with an average of 60.8 ± 6.1 ng/L (Fig. 1 and Table S1, SI). The highest PAH concentration in the coastal water of Dalian appeared at the lowest tide, while the lowest concentration was at the

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highest tide. The levels of PAHs in water in this study were comparable with those in water from the northern Gulf of Mexico (Σ43PAHs: 40 ng/L) (Adhikari et al., 2015),

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the Lake Superior (Σ21PAHs: 0.17-65 ng/L ), USA (Ruge et al., 2015), the

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north-eastern Mediterranean (Σ15PAHs: 2.4-25.9 ng/L) (Guitart et al., 2010), Incheon

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(Σ15PAHs: 8.34 ± 7.28 ng/L), South Korea (Kim and Chae, 2016), the Yangtze River

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of China (Σ16PAHs: 36 ng/L in summer; 41 ng/L in winter) (Lin et al., 2018), and the

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East China Sea (Σ15PAHs: 54 ng/L) (Ya et al., 2017), but one order of magnitude lower than those in the Daliao River (Σ16PAHs: 748 ng/L), China (Zheng et al., 2016) ,

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the Gomti River (Σ16PAHs: 1033-19940 ng/L), India (Malik et al., 2011), and Hai River estuary (Σ15PAHs: 1065 ng/L), China (Yan et al., 2016). 3.1.2 OPEs in seawater

The total concentrations of OPEs (Σ9OPEs) in the dissolved phase were in the range of 21.6-61.5 ng/L with an average of 39.7 ± 11.2 ng/L (Fig. 1). The highest concentration was measured at 11:00 on September 21st 2017, and the lowest concentration was at 17:00 on September 22nd 2017. The mean concentration of ΣOPEs was comparable with that in water from a freshwater lake of China (Σ12OPEs: 73.9 ng/L) (Xing et al., 2018), the Bohai and Yellow Seas (Σ7OPEs: 22.5 ng/L) 8

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(Zhong et al., 2017), a rural lake of Germany (Σ5OPEs: 60 ng/L) (Regnery and Puttmann, 2010), the Great Lakes (Σ6OPEs: 7.3-96 ng/L) (Venier et al., 2014), and the Lake Michigan Tributaries (Σ15OPEs: 20-54 ng/L) (Guo et al., 2017), but one order of magnitude lower than those in the North Sea, German (Σ12OPEs: 58.3-1092 ng/L) (Bollmann et al., 2012), rivers (Σ11OPEs: 300 ng/L) around the Bohai Sea,

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China (Wang et al., 2015), the Besòs River, Spain (Σ10OPEs: N.D.-7200 ng/L) (Cristale et al., 2013), and drinking water in Korea (Σ10OPEs: 140 ng/L) (Lee et al.,

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3.1.3 PAHs and OPEs in the air

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2016).

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Σ9PAHs in the gaseous phase were 30.6 and 16.3 ng/m3 on September 18th and

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September 25th, 2017, respectively (Table S2, SI). PHE was the dominant congener,

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accounting for 49.6% of the total concentration, followed by PYR (22.8%) and FLA (15.3%). The results were compared with a previous study in the coastal area of

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Dalian, China (15 PAHs, 27 ± 15 ng/m3) (Wu et al., 2019), the Lower Great Lakes, USA (15 PAHs, 2.1-76.4 ng/m3) (McDonough et al., 2014), the eastern India (Σ14PAHs: 17.1 ± 8.6 ng/m3) (Ray et al., 2017) and the Chaohu Lake, China (Σ16PAHs: 60.9 ± 46.3 ng/m3) (Qin et al., 2013). Σ9OPEs in the gaseous phase were 0.36 and 0.21 ng/m3 for these two samples, respectively. TCEP accounted 49.8% for the total concentration, followed by TCIPP (34.4%) and TNBP (8.7%). The air concentrations of OPEs were comparable with those detected in the air from the Bohai Sea and the Yellow Sea (Σ9OPEs: 0.1-0.75 ng/m3, particle and gas phases,) (Li et al., 2018) and the Mediterranean and Black Seas (Σ14OPEs: 0.4-6.0 ng/m3, particle 9

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and gas phases) (Castro-Jimenez et al., 2014). 3.2 Compositions of PAHs and OPEs 3.2.1 Composition of PAHs PHE was the most abundant congener in the dissolved phase, followed by FLU, ACE, and PYR, which contributed 35.3%, 21.7%, 10.4%, and 8.7% to the total PAHs,

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respectively (Fig. 1 and Fig. S1, SI). This congener profile is similar to those in water dissolved phase of Daliao River (PHE>FLU>ACE) (Zheng et al., 2016), coastal water

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of Korea (PHE>FLU) (Kim and Chae, 2016), and water of the Great Lakes

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(PHE>FLU>ACE) (Venier et al., 2014). This result can be attributed to the relatively

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low Kow (logKow: 3.4-4.6) of 3-ring PAHs, which made them more soluble in the

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water dissolved phase (McDonough et al., 2014). Liu et al. (Liu et al., 2013) also

oil spill accident.

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discovered that LMW PAHs accounted for 92% of PAHs in seawater samples after an

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3.2.2 Composition of OPEs

TCIPP was the predominant OPE congener in the dissolved phase, followed by TCEP, TBOEP, and TDCIPP with the contributions of 51.8%, 24.1%, 10.0%, and 6.3%, respectively (Fig. 1 and Fig. S1, SI). Similar OPE compositions were also observed in the water from Taihu Lake (TCIPP>TCEP>TBOEP) (Wang et al., 2018a), the Bohai and Yellow Seas, China (TCIPP>TCEP) (Zhong et al., 2017), and the Lake Michigan Tributaries (TCIPP>TCEP) (Guo et al., 2017). TCIPP, TCEP, and TDCIPP are three chlorinated OPEs, which are usually used as flame retardants, and their mass productions make them the most commonly used OPEs. TBOEP is usually used as a 10

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plasticizer in floor wax or in rubber products. Those OPEs are unanimously recalcitrant and persistent in environment due to their hydrophobic and toxic properties (Marklund et al., 2003; Reemtsma et al., 2008). 3.3 Tidal variability of PAHs and OPEs Diurnal variations were observed for both PAH and OPE concentrations in the

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coastal seawater, but with different trends, which may due to their different sources. The PAH concentrations showed a contrary diurnal variation with tide, which peaked

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at the low tides, and decreased with the flood tides. Σ9PAHs (r=-0.65, p<0.01), ANT

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(r=-0.64, p<0.01), and PHE (r=-0.51, p<0.05) showed significant negative

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correlations with tidal height, respectively (Table S3, SI). The diurnal variation of

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PAHs in this study followed the similar tidal variability to those of PAHs and PCBs in

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the Seine Estuary, where the lowest dissolved PAH and PCB concentrations in surface water were measured at the highest salinity period with the maximal tidal

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influence (Cailleaud et al., 2009). First reason for this pattern is that ebb or flow of tide makes a reciprocating motion for contaminant carried by water, which can accelerate their diffusions. The laminar flow in the tidal zone produced by the dynamic action has a significant influence on the transport of contaminant in the coastal water (Wu et al., 2017). Second reason is that tidal mixing process introduces a large amount of fresh offshore seawater with low level of PAHs into coastal water body, which can dilute the level of PAHs in the coastal water. Furthermore, offshore seawater with relatively high salinity may also cause “salting-out” effect, which can decrease the water solubility of a hydrophobic compound (Tremblay et al., 2005). 11

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The OPEs concentrations in the coastal water also varied over time, but no significant correlation was found between tidal heights and OPE concentrations,

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which may due to the occasional discharges of OPEs in the adjacent area. Compared

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with PAHs, OPE concentrations were generally lower and more strongly affected by

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irregular discharges from point sources, thus the instable OPE sources disrupted the

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tidal variability and made the correlation insignificant. This was proved by the

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significant negative correlations between ΣOPEs and water DO (r=-0.61, p<0.01) or pH (r=-0.62, p<0.01), and positive correlation between ΣOPEs and water DN

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(r=0.54, p<0.05) (Table S4, SI). 3.4 Source apportionments

3.4.1 Source apportionment of PAHs Fig. 2 shows the loading and score plots of PAHs in the water dissolved phase. The principal component 1(PC1), explaining 32.8% of the total variance, was dominated by CHR, PYR, ANT and PHE. Significant positive correlations were detected between PHE and ANT (r=0.77, p<0.05) as well as PYR and CHR (r=0.84, p<0.05) (Table S3, SI), implying similar sources of these PAH congeners. PC2, contributing 29.4%, was mainly influenced by BaA, ACY and FLA. Significant 12

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correlations of BaA and ACY (r=0.65, p<0.01) and BaA and FLA (r=0.71, p<0.01) were discovered, suggesting similar sources. However, samples distributed intensively in the score plot (Fig. 2b), indicating that the input sources of PAHs were stable during the sampling period. A 17 (samples) × 9 (PAH congeners) data set was input into the US EPA PMF

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5.0 model to estimate the source contributions of PAHs. A three-factor optimum solution with 20 runs was adopted after compared with solutions of 4 to 6 potential

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factors. As shown in Fig. 3, Factor 1, contributing 37.8% of the total PAHs, was

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mostly explained by the variation of ACY and ACE. ACY and ACE are typical

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markers of petrogenic sources, such as crude oil, and mainly caused by petroleum

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spill (Liu et al., 2009; Ghosal et al., 2013; Yu et al., 2015). This was consistent with

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the fact that the sampling site located near a small yacht harbor, thus the oil spill of yachts could be an important source of PAHs. Factor 2, contributing 32.5% of the

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total PAHs, was characterized by CHR, PYR, BaA and FLA. Previous studies suggested that PYR, FLA (Zhang et al., 2012), CHR and BaA (Wang et al., 2018b) are mainly derived from coal combustion. Therefore, Factor 2 indicated the coal combustion source. Factor 3, accounted for 29.7% of the total PAHs, was strongly dominated by FLU, which has been suggested to represent the gasoline vehicle exhaust (Ravindra et al., 2008; Gong et al., 2018). Therefore, Factor 3 might represent petroleum combustion, such as vehicle or ship exhausts (Chen et al., 2016; Zeng et al., 2018). However, because this source apportionment did not include 5-6-ring PAHs and conducted with limited sample size, results of the PMF analysis should be treated 13

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with caution. 3.4.2 Source apportionment of OPEs The loading and score plots of OPEs are shown in Fig. 2. The first and second principal components accounted for 34.4% and 21.1% of the variation, respectively. OPEs are frequently used as additives rather than chemically bonded to products,

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resulting in simple releases via volatilization and leaching during their lifetimes (Wei et al., 2015). Positive correlations of TNBP versus TDCIPP (r=0.64, p<0.01) and

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TNBP versus TMPP (r=0.52 p<0.05) were observed (Table S4, SI), suggesting that

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those three OPEs in the coastal water may originate from similar sources or undergo

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similar environmental processes. TDCIPP is frequently used as flame retardant, while

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TNBP and TMPP are primarily employed as plasticizers (Andresen et al., 2004).

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Furthermore, the n-octanol/water partition coefficients (log Kow) of TDCIPP (3.8), TNBP (4), and TMPP (5.11) are comparable (van der Veen and de Boer, 2012),

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implying these three contaminants may share similar transport or partition process. TCIPP and TCEP, which are generally applied as flame retardant, plasticizer, and lacquer, located close to each other in the loading plot (Fig. 2c). Seawater samples distributed dispersively in the score plot (Fig. 2d), suggesting OPEs in those samples may originate from multiple sources. 3.5 Fugacity fraction variation of PAHs and OPEs Fugacity fraction (ff) was calculated in order to determine the equilibrium statuses and gaseous exchange directions of PAHs and OPEs between water and air interface. The ff was calculated based on the PAH or OPE concentrations in the air 14

Journal Pre-proof (Ca, ng/m3) and water (Cw, ng/m3), air temperature (T, K), gas constant (R, Pa m3/(K mol)), and temperature-corrected Henry's law constant (H, Pa m3/mol) (van der Veen and de Boer, 2012; Wei et al., 2015) as follow: ff = fw/(fw+fa) = Cw/(Cw+Ca RT/H)

Eq(1)

Theoretically, an ff of 0.5 indicates that the chemical is equilibrium between air

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and water. However, due to the uncertainties, 0.26
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The ff values of PAHs and OPEs are shown in Fig. 4. The ff values of TEHP and

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TMPP were not calculated because their low concentrations in the air gaseous phase.

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Generally, ACY, ACE, FLU, PHE, and ANT displayed net volatilizations from water

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into air, while FLA, PYR, BaA, and CHR reached equilibrium between air and water.

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Our previous study (Wu et al., 2019) also suggested that the exchange trends of 3-ring PAHs were net volatilization from the coastal water into the air in Dalian, while the

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exchange trends of 4-ring PAHs varied with seasons and sampling sites. Although the ff values of PAHs varied slightly with tidal circle, this fluctuation was not strong enough to change their exchange directions between air and water in the study area. For OPEs, TCEP and TPHP were volatilized from the water into the air; TNBP and TDCIPP were deposited from the air into the water; while EHDPP reached equilibrium between air and water. However, the exchange trend of TCIPP varied with sampling time from deposition to equilibrium, which suggested that the air-water exchange of TCIPP in the coastal area was significantly impacted by the occasional discharges of TCIPP. Our previous study (Wang et al., 2018c) on OPEs also found 15

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that TCEP and TPHP were net volatilized from the water into the air with TDCIPP deposited from the air into the water, whereas TNBP and TCIPP varied with sampling season and location. The slight fluctuation of ff values of OPEs was also not strong enough to alter the air-water exchange directions of most congeners, except for TCIPP. This implied that the joint effect of tide and occasional discharge may

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sometimes change the air-water exchange trend of TCIPP in the coastal area. Here, we focused only on the gaseous exchange of PAHs and OPEs between air and

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the bulk depositions of PAHs and OPEs.

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seawater interface without considering particle deposition, which may underestimate

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4. Conclusion

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This study investigated the concentrations, tidal variablility, and air-water exchanges of PAHs and OPEs in a typical coastal area of Dalian, China. Obvious tidal

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variability was observed for PAHs in the coastal water, but not for OPEs, which may be attributed to the influence of occasional discharges of OPEs. Source apportionments suggested that PAHs in the coastal water were mainly from oil spill, traffic emission, and combustion, while OPEs came from multiple sources including wastewater discharge. The fugacity fractions of PAHs and OPEs varied slightly due to the influences of tide or source, but these influences were not strong enough to alter the air-water exchange directions of PAHs and most OPEs. However, the exchange of TCIPP was significantly impacted by the combined influence of both occasional discharge and tidal effect. Nevertheless, further investigations are still needed for the 16

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influence of tide on the environmental fates of chemicals in the coastal areas.

Appendix A. Supplementary data Table S1-S4, Figure S1, and S1. Supplementary data associated with this article

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can be found, in the online version.

Acknowledgements

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This study was supported by the Key Laboratory of Coastal Environmental

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Processes and Ecological Remediation, YICCAS (No. 2018KFJJ06), the Open Project

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of State Key Laboratory of Urban Water Resources and Environment (Grant No

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Reference

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(DUT19LK43).

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QA201945), and the Fundamental Research Funds for the Central Universities, China

Adhikari, P.L., Maiti, K., Overton, E.B., 2015. Vertical fluxes of polycyclic aromatic hydrocarbons in the northern Gulf of Mexico. Mar Chem 168, 60-68. Andresen, J.A., Grundmann, A., Bester, K., 2004. Organophosphorus flame retardants and plasticisers in surface waters. Sci Total Environ 332, 155-166. Bollmann, U.E., Moler, A., Xie, Z.Y., Ebinghaus, R., Einax, J.W., 2012. Occurrence and fate of organophosphorus flame retardants and plasticizers in coastal and marine surface waters. Water Res 46, 531-538. Bruhn, R., Lakaschus, S., McLachlan, M.S., 2003. Air/sea gas exchange of PCBs in the southern Baltic Sea. Atmos Environ 37, 3445-3454. Cailleaud, K., Forget-Leray, J., Peuhiet, L., LeMenach, K., Souissi, S., Budzinski, H., 2009. Tidal influence on the distribution of hydrophobic organic contaminants in the Seine Estuary and biomarker responses on the copepod Eurytemora affinis. Environ Pollut 157, 64-71. Cao, S.X., Zeng, X.Y., Song, H., Li, H.R., Yu, Z.Q., Sheng, G.Y., Fu, J.M., 2012. Levels and distributions of organophosphate flame retardants and plasticizers in sediment from Taihu Lake, China. Environ. Toxicol. Chem. 31, 1478-1484. Castro-Jimenez, J., Berrojalbiz, N., Pizarro, M., Dachs, J., 2014. Organophosphate Ester (OPE) Flame 17

Journal Pre-proof Retardants and Plasticizers in the Open Mediterranean and Black Seas Atmosphere. Environ Sci Technol 48, 3203-3209. Chen, Y.J., Lin, T., Tang, J.H., Xie, Z.Y., Tian, C.G., Li, J., Zhang, G., 2016. Exchange of polycyclic aromatic hydrocarbons across the air-water interface in the Bohai and Yellow Seas. Atmos Environ 141, 153-160. Cristale, J., Vazquez, A.G., Barata, C., Lacorte, S., 2013. Priority and emerging flame retardants in rivers: Occurrence in water and sediment, Daphnia magna toxicity and risk assessment. Environ Int 59, 232-243. Ghosal, D., Dutta, A., Chakraborty, J., Basu, S., Dutta, T.K., 2013. Characterization of the metabolic pathway involved in assimilation of acenaphthene in Acinetobacter sp strain AGAT-W. Res Microbiol 164, 155-163. Gong, P., Wang, X.P., Sheng, J.J., Wang, H.L., Yuan, X.H., He, Y.Q., Qian, Y., Yao, T.D., 2018.

of

Seasonal variations and sources of atmospheric polycyclic aromatic hydrocarbons and organochlorine compounds in a high-altitude city: Evidence from four-year observations. Environ

ro

Pollut 233, 1188-1197.

Guitart, C., Garcia-Flor, N., Miquel, J.C., Fowler, S.W., Albaiges, J., 2010. Effect of the accumulation

-p

of polycyclic aromatic hydrocarbons in the sea surface microlayer on their coastal air-sea exchanges. J Marine Syst 79, 210-217.

re

Guo, J.H., Romanak, K., Westenbroek, S., Hites, R.A., Venier, M., 2017. Current-Use Flame Retardants in the Water of Lake Michigan Tributaries. Environ Sci Technol 51, 9960-9969. Jurado, E., Jaward, F., Lohmann, R., Jones, K.C., Simo, R., Dachs, J., 2005. Wet deposition of

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persistent organic pollutants to the global oceans. Environ Sci Technol 39, 2426-2435. Kawagoshi, Y., Nakamura, S., Fukunaga, I., 2002. Degradation of organophosphoric esters in leachate from a sea-based solid waste disposal site. Chemosphere 48, 219-225.

na

Kim, S.K., Chae, D.H., 2016. Seasonal variation in diffusive exchange of polycyclic aromatic hydrocarbons across the air-seawater interface in coastal urban area. Mar Pollut Bull 109, 221-229.

Jo ur

Lee, S., Jeong, W., Kannan, K., Moon, H.B., 2016. Occurrence and exposure assessment of organophosphate flame retardants (OPFRs) through the consumption of drinking water in Korea. Water Res 103, 182-188.

Li, C.H., Zhou, H.W., Wong, Y.S., Tam, N.F.Y., 2009. Vertical distribution and anaerobic biodegradation of polycyclic aromatic hydrocarbons in mangrove sediments in Hong Kong, South China. Sci Total Environ 407, 5772-5779. Li, J., Tang, J.H., Mi, W.Y., Tian, C.G., Emeis, K.C., Ebinghaus, R., Xie, Z.Y., 2018. Spatial Distribution and Seasonal Variation of Organophosphate Esters in Air above the Bohai and Yellow Seas, China. Environ Sci Technol 52, 89-97. Lin, L., Dong, L., Meng, X.Y., Li, Q.Y., Huang, Z., Li, C., Li, R., Yang, W.J., Crittenden, J., 2018. Distribution and sources of polycyclic aromatic hydrocarbons and phthalic acid esters in water and surface sediment from the Three Gorges Reservoir. J Environ Sci-China 69, 271-280. Liu, F., Hu, S., Guo, X.J., Niu, L.X., Cai, H.Y., Yang, Q.S., 2018. Impacts of estuarine mixing on vertical dispersion of polycyclic aromatic hydrocarbons (PAHs) in a tide-dominated estuary. Mar Pollut Bull 131, 276-283. Liu, F., Niu, L.X., Chen, H., Li, P., Tian, F., Yang, Q.S., 2017. Seasonal changes of polycyclic aromatic hydrocarbons in response to hydrology and anthropogenic activities in the Pearl River estuary, China. Mar Pollut Bull 117, 255-263. 18

Journal Pre-proof Liu, F., Yang, Q.S., Hu, Y.J., Du, H.H., Yuan, F., 2014. Distribution and transportation of polycyclic aromatic hydrocarbons (PAHs) at the Humen river mouth in the Pearl River delta and their influencing factors. Mar Pollut Bull 84, 401-410. Liu, X.J., Jia, H.L., Wang, L., Qi, H., Ma, W.L., Hong, W.J., Guo, J.G., Yang, M., Sun, Y.Q., Li, Y.F., 2013. Characterization of polycyclic aromatic hydrocarbons in concurrently monitored surface seawater and sediment along Dalian coast after oil spill. Ecotox Environ Safe 90, 151-156. Liu, Y., Chen, L., Huang, Q.H., Li, W.Y., Tang, Y.J., Zhao, J.F., 2009. Source apportionment of polycyclic aromatic hydrocarbons (PAHs) in surface sediments of the Huangpu River, Shanghai, China. Sci Total Environ 407, 2931-2938. Lohmann, R., Dapsis, M., Morgan, E.J., Dekany, V., Luey, P.J., 2011. Determining Air-Water Exchange, Spatial and Temporal Trends of Freely Dissolved PAHs in an Urban Estuary Using Passive Polyethylene Samplers. Environ Sci Technol 45, 2655-2662.

of

Malik, A., Verma, P., Singh, A.K., Singh, K.P., 2011. Distribution of polycyclic aromatic hydrocarbons in water and bed sediments of the Gomti River, India. Environ Monit Assess 172,

ro

529-545.

Marklund, A., Andersson, B., Haglund, P., 2003. Screening of organophosphorus compounds and their

-p

distribution in various indoor environments. Chemosphere 53, 1137-1146. Maskaoui, K., Zhou, J.L., Hong, H.S., Zhang, Z.L., 2002. Contamination by polycyclic aromatic

re

hydrocarbons in the Jiulong River Estuary and Western Xiamen Sea, China. Environ Pollut 118, 109-122.

McDonough, C.A., Khairy, M.A., Muir, D.C.G., Lohmann, R., 2014. Significance of Population Technol 48, 7789-7797.

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Centers As Sources of Gaseous and Dissolved PAHs in the Lower Great Lakes. Environ Sci Meijer, S.N., Shoeib, M., Jantunen, L.M.M., Jones, K.C., Harner, T., 2003. Air-soil exchange of

na

organochlorine pesticides in agricultural soils. 1. Field measurements using a novel in situ sampling device. Environ Sci Technol 37, 1292-1299.

Jo ur

Moller, A., Sturm, R., Xie, Z.Y., Cai, M.H., He, J.F., Ebinghaus, R., 2012. Organophosphorus Flame Retardants and Plasticizers in Airborne Particles over the Northern Pacific and Indian Ocean toward the Polar Regions: Evidence for Global Occurrence. Environ Sci Technol 46, 3127-3134. Motelay-Massei, A., Garban, B., Phagne-Larcher, K., Chevreuil, M., Ollivon, D., 2006. Mass balance for polycyclic aromatic hydrocarbons in the urban watershed of Le Havre (France): Transport and fate of PAHs from the atmosphere to the outlet. Water Res 40, 1995-2006. Ou, Y., 2011. Developments of organic phosphorus flame retardant industry in China. Chemical Industry and Engineering Progress 30, 210-215. Qin, N., He, W., Kong, X.Z., Liu, W.X., He, Q.S., Yang, B., Ouyang, H.L., Wang, Q.M., Xu, F.L., 2013. Atmospheric partitioning and the air-water exchange of polycyclic aromatic hydrocarbons in a large shallow Chinese lake (Lake Chaohu). Chemosphere 93, 1685-1693. Ravindra, K., Sokhi, R., Van Grieken, R., 2008. Atmospheric polycyclic aromatic hydrocarbons: Source attribution, emission factors and regulation. Atmos Environ 42, 2895-2921. Ray, D., Chatterjee, A., Majumdar, D., Ghosh, S.K., Raha, S., 2017. Polycyclic aromatic hydrocarbons over a tropical urban and a high altitude Himalayan Station in India: Temporal variation and source apportionment. Atmos Res 197, 331-341. Reemtsma, T., Quintana, J.B., Rodil, R., Garcia-Lopez, M., Rodriguez, I., 2008. Organophosphorus flame retardants and plasticizers in water and air I. Occurrence and fate. Trac-Trend Anal Chem 27, 19

Journal Pre-proof 727-737. Regnery, J., Puttmann, W., 2010. Occurrence and fate of organophosphorus flame retardants and plasticizers in urban and remote surface waters in Germany. Water Res 44, 4097-4104. Ruge, Z., Muir, D., Helm, P., Lohmann, R., 2015. Concentrations, Trends, and Air-Water Exchange of PAHs and PBDEs Derived from Passive Samplers in Lake Superior in 2011. Environ Sci Technol 49, 13777-13786. Sankoda, K., Kuribayashi, T., Nomiyama, K., Shinohara, R., 2013. Occurrence and Source of Chlorinated Polycyclic Aromatic Hydrocarbons (CI-PAHs) in Tidal Flats of the Ariake Bay, Japan. Environ Sci Technol 47, 7037-7044. Tremblay, L., Kohl, S.D., Rice, J.A., Gagne, J.P., 2005. Effects of temperature, salinity, and dissolved humic substances on the sorption of polycyclic aromatic hydrocarbons to estuarine particles. Mar Chem 96, 21-34.

of

van der Veen, I., de Boer, J., 2012. Phosphorus flame retardants: Properties, production, environmental occurrence, toxicity and analysis. Chemosphere 88, 1119-1153.

ro

Venier, M., Dove, A., Romanak, K., Backus, S., Hites, R., 2014. Flame Retardants and Legacy Chemicals in Great Lakes' Water. Environ Sci Technol 48, 9563-9572.

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Wang, R.M., Tang, J.H., Xie, Z.Y., Mi, W.Y., Chen, Y.J., Wolschke, H., Tian, C.G., Pan, X.H., Luo, Y.M., Ebinghaus, R., 2015. Occurrence and spatial distribution of organophosphate ester flame

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retardants and plasticizers in 40 rivers draining into the Bohai Sea, north China. Environ Pollut 198, 172-178.

Wang, X.L., Zhu, L.Y., Zhong, W.J., Yang, L.P., 2018a. Partition and source identification of

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organophosphate esters in the water and sediment of Taihu Lake, China. J Hazard Mater 360, 43-50.

Wang, X.T., Hu, B.P., Cheng, H.X., Jia, H.H., Zhou, Y., 2018b. Spatial variations, source

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apportionment and potential ecological risks of polycyclic aromatic hydrocarbons and synthetic musks in river sediments in Shanghai, China. Chemosphere 193, 108-117.

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Wang, Y., Wu, X.W., Zhang, Q.N., Zhao, H.X., Hou, M.M., Xie, Q., Chen, J.W., 2018c. Occurrence, distribution, and air-water exchange of organophosphorus flame retardants in a typical coastal area of China. Chemosphere 211, 335-344. Wei, G.L., Li, D.Q., Zhuo, M.N., Liao, Y.S., Xie, Z.Y., Guo, T.L., Li, J.J., Zhang, S.Y., Liang, Z.Q., 2015. Organophosphorus flame retardants and plasticizers: Sources, occurrence, toxicity and human exposure. Environ Pollut 196, 29-46. Wu, X.W., Wang, Y., Zhang, Q.N., Zhao, H.X., Yang, Y., Zhang, Y.W., Xie, Q., Chen, J.W., 2019. Seasonal variation, air-water exchange, and multivariate source apportionment of polycyclic aromatic hydrocarbons in the coastal area of Dalian, China. Environ Pollut 244, 405-413. Wu, Y.L., Wang, X.H., Li, Y.Y., Ya, M.L., Luo, H., Hong, H.S., 2017. Polybrominated diphenyl ethers, organochlorine pesticides, and polycyclic aromatic hydrocarbons in water from the Jiulong River Estuary, China: levels, distributions, influencing factors, and risk assessment. Environ Sci Pollut R 24, 8933-8945. Xing, L.Q., Zhang, Q., Sun, X., Zhu, H.X., Zhang, S.H., Xu, H.Z., 2018. Occurrence, distribution and risk assessment of organophosphate esters in surface water and sediment from a shallow freshwater Lake, China. Sci Total Environ 636, 632-640. Ya, M.L., Wang, X.H., Wu, Y.L., Li, Y.Y., Yan, J.M., Fang, C., Zhao, Y.Y., Qian, R.R., Lin, X.L., 2017. Seasonal Variation of Terrigenous Polycyclic Aromatic Hydrocarbons along the Marginal 20

Journal Pre-proof Seas of China: Input, Phase Partitioning, and Ocean-Current Transport. Environ Sci Technol 51, 9072-9079. Ya, M.L., Wang, X.H., Wu, Y.L., Ye, C.X., Li, Y.Y., 2014. Enrichment and partitioning of polycyclic aromatic hydrocarbons in the sea surface microlayer and subsurface water along the coast of Xiamen Island, China. Mar Pollut Bull 78, 110-117. Yan, J.X., Liu, J.L., Shi, X., You, X.G., Cao, Z.G., 2016. Polycyclic aromatic hydrocarbons (PAHs) in water from three estuaries of China: Distribution, seasonal variations and ecological risk assessment. Mar Pollut Bull 109, 471-479. Yu, W.W., Liu, R.M., Wang, J.W., Xu, F., Shen, Z.Y., 2015. Source apportionment of PAHs in surface sediments using positive matrix factorization combined with GIS for the estuarine area of the Yangtze River, China. Chemosphere 134, 263-271. Zeng, Q.F., Jeppesen, E., Gu, X.H., Mao, Z.G., Chen, H.H., 2018. Distribution, fate and risk

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assessment of PAHs in water and sediments from an aquaculture- and shipping-impacted subtropical lake, China. Chemosphere 201, 612-620.

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Zhang, K., Liang, B., Wang, J.Z., Guan, Y.F., Zeng, E.Y., 2012. Polycyclic aromatic hydrocarbons in upstream riverine runoff of the Pearl River Delta, China: An assessment of regional input sources.

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Environ Pollut 167, 78-84.

Zhang, Z.L., Hong, H.S., Zhou, J.L., Yu, G., 2004. Phase association of polycyclic aromatic

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hydrocarbons in the Minjiang River Estuary, China. Sci Total Environ 323, 71-86. Zheng, B.H., Wang, L.P., Lei, K., Nan, B.X., 2016. Distribution and ecological risk assessment of polycyclic aromatic hydrocarbons in water, suspended particulate matter and sediment from Daliao

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River estuary and the adjacent area, China. Chemosphere 149, 91-100. Zhong, M.Y., Tang, J.H., Mi, L.J., Li, F., Wang, R.M., Huang, G.P., Wu, H.F., 2017. Occurrence and spatial distribution of organophosphorus flame retardants and plasticizers in the Bohai and Yellow

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Fig. 1 Tidal height and concentrations of PAHs and OPEs in the coastal water.

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Fig. 2 Two-dimensional principal component for PAHs (loading plot (a) and score

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Fig. 3 Source apportionments of PAHs by PMF model in Dalian.

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Fig. 4 Air-water fugacity fractions of PAHs and OPEs.

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Graphical abstract

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Highlights:  Significant tidal variability was found for PAHs in coastal water but not for OPEs.  Tide was not strong enough to alert the air-water exchange trends of PAHs or OPEs.

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 Combined effect of source and tide can alert air-water exchange direction of

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