Fossil Fuel-Derived Polycyclic Aromatic Hydrocarbons in the Taiwan

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Environmental Processes

Fossil Fuel-Derived Polycyclic Aromatic Hydrocarbons in the Taiwan Strait, China, and Fluxes across the Air-Water Interface Miaolei Ya, Li Xu, Yu-Ling Wu, Yongyu Li, Songhe Zhao, and Xin-Hong Wang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b01331 • Publication Date (Web): 01 Jun 2018 Downloaded from http://pubs.acs.org on June 1, 2018

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Environmental Science & Technology

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Fossil Fuel-Derived Polycyclic Aromatic Hydrocarbons in the Taiwan Strait, China,

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and Fluxes across the Air-Water Interface

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Miaolei Ya1, 2, Li Xu2 *, Yuling Wu1, Yongyu Li1, Songhe Zhao1 and Xinhong Wang1 *.

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1

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Ecology, Xiamen University, Xiamen 361102, China.

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2

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and Geophysics, Woods Hole Oceanographic Institution, Woods Hole, Massachusetts

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02543, United States.

State Key Laboratory of Marine Environmental Science, College of the Environment & National Ocean Sciences Accelerator Mass Spectrometry Facility, Department of Geology

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ABSTRACT

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Based on the application of compound-specific radiocarbon analysis (CSRA) and air-

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water exchange models, the contributions of fossil fuel and biomass burning derived

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polycyclic aromatic hydrocarbons (PAHs) as well as their air-water transport were

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elucidated. The results showed that fossil fuel-derived PAHs (an average contribution of

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89%) present the net volatilization process at the air-water interface of the Taiwan Strait in

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summer. Net volatile fluxes of the dominant fluorene and phenanthrene (>58% of the total

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PAHs) were 27±2.8 μg m−2∙day−1, significantly higher than the dry deposition fluxes

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(average 0.43 μg m−2∙day−1). The Δ14C contents of selected PAHs (fluorene, phenanthrene

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plus anthracene, fluoranthene, and pyrene) determined by CSRA in the dissolved seawater

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ranged from −997±4‰ to −873±6‰, indicating 89−100% (95±4%) of PAHs were

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contributed by fossil fuels. The South China Sea warm current originating from the

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southwest China in summer (98%) and the Min-Zhe coastal current originating from the

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north China in winter (97%) input more fossil fuel PAHs than the Jiulong River estuary

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(90%) and Xiamen harbor water (93%). The more radioactive decayed 14C of fluoranthene

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(a 4- ring PAH) than phenanthrene and anthracene (3- ring PAHs) represented a greater

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fossil fuel contribution to the former in dissolved seawater.

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ACS Paragon Plus Environment


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1. INTRODUCTION

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The origin of polycyclic aromatic hydrocarbons (PAHs) and air-water interface

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transport are important environmental processes affecting the exist and fate of PAHs in

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marine environment.1, 2 Generally, sources of PAHs to the coastal ocean include, but are

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not limited to, the incomplete combustion and pyrolysis of fossil fuels (e.g., coal, petroleum)

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or modern biomass (e.g., wood, straw and grass and other C3 and C4 plants) and the direct

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release of petroleum products.3, 4 The research on the origin of PAHs is to find out which

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type of human activity is the most important way for PAHs to import into the offshore

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ocean. After the land-emitting and ocean-based anthropogenic PAHs are transmitted to

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marine environment, they are transported across regional seas under the action of coastal

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water masses.5-7 Synchronously, the dynamic exchanges (atmospheric deposition and free

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exchange) of PAHs occur at the air-water interface.8, 9 Air-water transport of PAHs is

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proved to be a critical intermediate process that affects their migration to the global ocean

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and burial into the sedimentary environment.10-12 PAH species vary in their volatilization

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/deposition processes at the air-water interface.13 Typically, the air-water interface is a

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source of low molecular weight (LMW) PAHs and a sink of high molecular weight (HMW)

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PAHs in the coastal ocean.14,

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radiocarbon analysis (CSRA) and air-water exchange models to quantitatively study the

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sources (fossil fuel vs. biomass burning) of PAHs in surface seawater and interface

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transport (deposition or volatilization), respectively, in the coastal water masses of the

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Taiwan Strait (TWS), connecting the South China Sea (SCS) and the East China Sea (ECS).

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Traditionally, the source identifications of PAHs were widely studied using the isomer

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diagnostic ratios and receptor models such as principal component analysis (PCA), and

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positive matrix factorization, etc.1, 16-18 However, variations in combustion conditions and

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environmental degradation processes occurring during the transport of PAH isomers from

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their sources to receptors 19 undermines the application of diagnostic ratios and statistical

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methods as reliable source apportionment tools.20 Therefore, the advent of CSRA based on

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the separation of individual compounds by preparative capillary gas chromatography

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(PCGC) augments isotope composition heterogeneity by the molecular level

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Therefore, we applied advanced compound-specific

21, 22

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C-

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Accelerator Mass Spectrometry (AMS) measurement.

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deconvoluting PAH sources in complex, heterogeneous matrices (soil, sedimentary and

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atmospheric environments).23-26 These studies have capitalized on the difference in

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radiocarbon contents between fossil fuel sources of PAHs, which contain no significant 2


This has proven useful for

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radiocarbon (Δ14C = −1000‰; Δ14C is expressed as a deviation from a known “oxalic acid”

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standard, which normalizes the radiocarbon content of a sample to the same δ13C (−25‰)

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and time point (1950)27, 28), and biomass burning sources, which contain modern levels of

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radiocarbon (such as wood burning with a Δ14C of +225‰ or leaves and annual grass

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burning with a Δ14C of +70‰).29 Using a radiocarbon isotopic mass balance approach, the

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relative contributions of these sources to PAHs presented in environmental matrixes can be

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apportioned.30 However, limited by the complexity and high demands of ultra-microscale

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(ca. 5.0−25 μg C) radiocarbon measurements of individual pure compounds,24, 31 CSRA

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method was used only in the identification of PAHs sources in the atmospheric, soil and

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sediment environments.23-26, 29, 30, 32 Up to now, this method has not been applied to the

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source identification of PAHs in the aqueous medium, limited by the large volume

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enrichment of trace PAHs (e.g., phenanthrene (Phen): 21−71 ng L−1 in our previous study33,

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34

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sedimentation and burial of PAHs.35, 36 Therefore, it is a vital and worthy research topic to

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extend the application of CSRA to assess the source characteristic of PAHs in the aquatic

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environment.

) in seawater. However, coastal seawater plays an important role in global transport,

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In this study, we focused on the CSRA application in four typical water masses in the

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TWS,37 which represent different PAH origins and also have differing effects on the

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transport of anthropogenic PAHs. In detail, Jiulong River diluted water (JRDW), which has

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the largest runoff discharge in summer, likely contains PAHs emitted from industrial and

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agricultural activities (e.g., sewage discharges and surface runoff) in surrounding rural and

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urban areas along the its drainage path (Figure 1).33, 38 Xiamen Western Harbor (XMWH),

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the biggest cargo terminal in the western TWS, can be assumed to contain fossil fuel-

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derived PAHs from the release and combustion of petroleum in the port (e.g., shipping

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activities) and from urban vehicle-derived emissions.34 Min-Zhe coastal current (MZCC),

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driven by the northeast monsoon in winter, could transport anthropogenic PAHs from the

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Yangtze River and Hangzhou Bay along the coast of the ECS.39 The South China Sea warm

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current (SCSWC), originating from the Southeast Asian coasts, represents the long-range

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transport of PAHs driven by the southwest monsoon in summer.5, 40 At present study, we

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determine the source (fossil fuel or biomass burning) and their contributions of PAHs by

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the measurement of the 14C of individual PAHs, based on CSRA in the above-mentioned

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water masses. Additionally, we also reveal the transport characteristic of fossil fuel versus

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biomass burning derived PAHs at the air-water interface by examining transport behaviors

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(volatilization or deposition and their fluxes) of PAHs in the TWS. This will not only 3



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provide scientific advice on the reduction of PAHs, but also have important implications

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for studying the migration and burial of PAHs in the vertical direction of the offshore

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seawater.

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2. MATERIALS AND METHODS

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2.1. Sampling of PAHs.

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Sampling for Concentration Analysis During the summer (June and July) of 2010,

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13 surface seawater (sites T1 to T13; ~1 m depth) and 3 aerosol samples (sites TW01,

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TW02, and TW03: TW01 covered site T1-T6 area; TW02 covered site T6-T10 area; TW03

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covered site T10-end area) were taken from the R/V Yanping 2 in the central TWS (solid

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black spots in Figure 1). Seawater sampling procedures have been described in our previous

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studies (see section 1 of the Supporting Information, SI),5, 41 glass-fiber filters (GF/Fs) and

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solid-phase extraction (SPE) cartridges were used to collect suspended particulate matters

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(SPM, particulate phase) and dissolved phase samples. Synchronously, aerosol sampling

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was carried out on the forecastle of the ship (upwind of the chimney, about 5 m high) using

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large flow air samplers (Thermo Electron Corporation) while sailing against the wind to

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avoid the contamination from the ship, such as chimney. Quartz filters (O.D. 100 mm,

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Advantec Corporation, pre-burned at 600 °C for 6 hours) and polyurethane foams (PUF,

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O.D. 65 mm, length 40 mm; Soxhlet extracted with dichloromethane for 24 h) were used

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to synchronously collect total suspended particulates (TSP, aerosol phase) and gas phase

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(PUF phase) samples at a flow speed of 360 L min−1. Finally, GF/Fs and SPE cartridges

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from seawater sampling and TSP and PUF from air sampling were hermetically stored at

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−20 °C until further analysis.

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Figure 1 Sampling areas in the TWS. The roman numerals from I to IV show the bulk seawater

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sampling locations where the Δ14C values of selected PAHs were measured. MZCC: Min-Zhe

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coastal current; SCSWC: South China Sea warm current; JRDW: Jiulong River diluted water;

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XMWH: Xiamen Western Harbor.

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Bulk Sampling for 14C Analysis To measure the Δ14C and δ13C values (relative to

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VPDB: Vienna Pee Dee Belemnite) of selected PAHs dissolved in TWS seawater, bulk

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seawater samples (5000–20000 L) were collected from the previously described four

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regional water masses, including the MZCC (I), SCSWC (II), JRDW (III) and XMWH (IV)

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(Figure 1). After initial filtration by GF/Fs (Whatman, O.D. 142 mm, 0.75 μm), dissolved

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organic matter (DOM, 1500 mL of XAD-2 resin) were used to extract the DOM to collect enough

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target dissolved PAHs from the seawater. Prior to use, XAD-2 resin was regenerated

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following standard procedures (see section 2 of the SI) and ultrasonically extracted with 5



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ethyl alcohol (all solvents from TEDIA Company, USA). The XAD-2 resin was then

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transferred into the glass columns mentioned above. Ethyl alcohol was removed from the

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gaps under a gentle stream of N2 (99.99%). Then, the XAD-2 columns were pre-extracted

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with 100 mL acetone, 100 mL dichloromethane (DCM) and 100 mL n-hexane at a slow

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flow speed (drop-by-drop) to remove contaminants in the matrix. Finally, plenty of

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ultrapure water (18.25 MΩ·cm, Millipore Company) was used to elute the residual solvents.

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After field sampling, used XAD-2 columns were stored at 4 °C until further pretreatment

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for CSRA steps.

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2.2. Concentration Analysis.

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The analytical procedures of PAHs concentrations for the GF/F and SPE seawater

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samples have been described in more detail in our previous research.33,

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pretreatment procedure for atmospheric TSP samples was the same as for the GF/F samples.

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For the atmospheric PUF samples, Soxhlet extraction was carried out for 24 h after spiking

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with five deuterated PAH surrogates. After rotary evaporation, the concentrated extract was

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cleaned using pretreated alumina/silica gel chromatography. Finally, the eluate was blown

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gently under a stream of N2 gas to 100 μL and stored at –20 °C before instrumental analysis.

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The detailed sample pretreatment methods and instrumental analysis by Agilent 6890

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Series gas chromatography–Agilent 5973 mass spectrometry are described in our previous

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papers.5,

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designated by the United States Environmental Protection Agency. Their abbreviations are

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shown in section 3 of the SI.

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2.3. 14C Preparation and Analysis.

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34, 38

The

The target compounds including 15 priority PAHs (excluding naphthalene)

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Based on a previous field investigation of individual PAH concentrations (e.g.,

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phenanthrene (Phen): 21−71 ng L−1) in the seawater of the TWS,33, 34 seawater sampling

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for several thousand liters is required to isolate adequate quantities of PAHs for 14C-AMS

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measurement.

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For the bulk seawater samples, residual seawater was drained from the XAD-2 column

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under a gentle stream of N2 and then discarded. Each XAD-2 column was eluted with 100

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mL methanol followed by 200 mL DCM at a slow flow rate (drop by drop). Residual

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solvents were removed from the XAD-2 under a gentle stream of N2. The mixed extracts

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were transferred into a 1 L separatory funnel, to which 150 mL of ultrapure water was

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added. After sufficient mixing and subsequent phase separation, 100 mL DCM was added

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3x for liquid/liquid extractions. DCM extracts were concentrated to near dryness and the

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solvent was exchanged with 5 mL n-hexane. After being concentrated to ~1 mL, extracts

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were fractionated in a glass column (i.d.15 mm) packed with 10 g of activated silica (0.063–

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0.20 mm mesh, activated at 450 °C for 4 h and deactivated at 130 °C for 12 h). The elution

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solvents used are as follows: fraction 1, 25 mL of hexane followed by 10 mL of DCM/n-

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hexane (v: v, 1:1); fraction 2, 40 mL of DCM/n-hexane. Fraction 2, containing the PAHs,

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was concentrated and filtered through a syringe-tip filter (Millipore 4 mm Millex-FH, pore

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size 0.2 µm). Finally, all extracts were combined and concentrated to ~1 mL. The PAH

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fraction was cleaned by partitioning with n-pentane/dimethylformamide, dried with

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anhydrous sodium sulfate,42 and reduced to the proper volume under a gentle stream of N2.

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PAHs were quantified using gas chromatography with a flame ionization detector (GC-

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FID).

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After purification, PCGC was used to isolate and prepare individual PAHs.21, 29, 43 The

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PCGC system consisted of an Agilent 6890 GC with a FID and a 7683 Series auto-injector,

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combined with a cooled injection system (CIS, Gerstel) and a preparative fraction collector

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(PFC, Gerstel) operated at 320 °C. Details of the PCGC method used for PAH isolation are

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listed in section 4 of the SI. Briefly, individual PAHs were isolated using two GC columns

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(30 m Agilent HP5 and 30 m Agilent DB17) in sequence and collected by the PFC 2x.

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Approximately 1% of the column eluate was diverted to the FID detector and the remaining

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99% of the sample was collected into glass U-tubes by the PFC. The individual PAHs

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including fluorene (Flu), Phen plus anthracene (An), fluoranthene (Fluo), and pyrene (Py)

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were collected into the independent glass U-tubes. Phen and An cannot be separated by

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DB17 chromatographic column, so we combined these two isomers together as one

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component for

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acenaphthylene (Ace), acenaphthene (Acen), benzo(a)anthracene (BaA), chrysene (Chry),

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and 5- and 6- ring PAHs (see section 3 of the SI) in the dissolved phase were too low to be

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performed for radiocarbon analysis. Trapped individual PAHs were transferred to 4 mL

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vials by rinsing with n-hexane 4x. Finally, the samples were passed through a silica gel

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column (4.0 cm × 0.5 cm i.d.) to remove column bleed from the PCGC. A small aliquot

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from each trap was taken to check the purity of the PAHs by GC-FID.

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C analysis. The abundances of the other detected PAHs such as

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Carbon isotope analysis was performed at the National Ocean Sciences Accelerator

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Mass Spectrometry (NOSAMS) facility at Woods Hole Oceanographic Institute. After the

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catalytic oxidation of the target PAHs, the generated CO2 was purified and quantified in

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the vacuum glass line. Finally, the CO2 was reduced to graphite and Δ14C measurements 7



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were made by AMS.44 Detailed procedures are shown elsewhere.29 The Δ14C values based

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on the sampling year were compared to the NBS oxalic acid I (NIST SRM 4990) standard

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after normalizing the radiocarbon content of a sample to the same δ13C (−25‰) and time

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point (1950).27, 45

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2.4. Quality Assurance and Quality Control.

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We strictly performed analytic procedures to ensure quality control. Field blanks of

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GF/Fs, quartz filters, SPE cartridges and PUF were taken to represent potential

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contamination during handling on the ship. Concentrations of target PAHs in the field blank

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samples were ~2 orders of magnitude less than in the seawater samples, which means we

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can largely ignore the PAH background from the ship and laboratory. For the water samples,

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the method recovery of the 15 PAHs ranged from 74% to 126% in the dissolved phase (SPE

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cartridges) and from 60% to 120% in the particulate phase (baked GF/Fs). For the

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atmospheric samples, the method recovery ranged from 78% to 105% in the aerosol

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samples (baked quartz filters) and from 72% to 113% in the gas samples (clean PUF). The

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method detection limits (MDLs) were calculated as the mean plus three times the standard

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deviation of the field blanks. For the water samples, MDLs of the 15 PAHs ranged from

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0.9–4.1 pg L−1 in the dissolved phase and from 0.8–3.5 pg L−1 in the particulate phase

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(considering an average sampling volume of 8 L). For the atmospheric samples, MDLs

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ranged from 0.7–2.4×102 pg m−3 in the aerosol phase and from 1.8–4.3×102 pg m−3 in the

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gas phase (considering the sampled volume of 400 m3).

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Prior to bulk seawater sampling, the XAD-2 resin was purified, in order, with acetone,

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DCM and n-hexane to remove contaminants in the matrix. The mass of PAHs remaining in

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the 100 mL of purified XAD-2 resin ranged from n. d. (not detected) to 25 ng. This showed

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that PAHs in the matrix were 2 orders of magnitude less abundant than in the field samples

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(>17 µg). PAHs in the solvents used in the pretreatment procedure were not detected and

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other impurities in the solvent or introduced during sample handling were removed during

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the PCGC step. Solvent from the final column bleed cleanup (less than 4 mL) was blew to

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dryness in a baked quartz tube before combustion. Therefore, to the trapped target PAHs,

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the pretreatment, PCGC collection and column bleed removal steps could also not cause

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additional carbon contamination. Studies have shown that the retention time deviation of

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the PCGC collection process is the main cause of isotope fractionation. Therefore, we

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strictly controlled the deviation of retention time during collection. And our previous work

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has shown no isotope fractionation of

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C during the PCGC collection step.29, 46 Finally, 8


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multiple PCGC blanks were combined to a concentration of 0.15 μmol C, and the fraction

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modern value of the combined blank was 0.29.

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3. RESULTS AND DISCUSSION

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3.1. PAHs in Water and Air of the TWS.

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The sum of the 15 PAH concentrations (Σ15PAHs) in the central TWS (sites T1 to T13,

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n=13) ranged from 53 to 79 (62±8.3) ng L−1 in the dissolved seawater (Table S1 and Figure

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S1), and from 2.5 to 5.6 (4.0±1.1) ng L−1 in the particulate phase in seawater (Table S2).

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As the dominant phase of PAHs (>91%, Figure S2), dissolved PAH concentrations showed

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an even distribution in the central TWS (Figures 2a and S3). The composition of PAHs also

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showed no significant difference among the different stations (p>0.05). Phen (28±4.0 ng

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L−1) followed by Flu (20±2.9 ng L−1) displayed the highest relative abundance with average

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contributions of 46% and 32% to Σ15PAHs across the TWS, respectively. Phen and Flu, the

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predominant components in ultrafine particles of petroleum combustion in the engines,47

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might derived from the emissions of vehicles and ships in surrounding cities and cruise

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routes in the TWS. The abundances of these two compounds were higher than found in the

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open southern Indian Ocean (Phen: 1.7 ng L−1; Flu: 1.4 ng L−1),6 close to the levels in the

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ECS (Phen: 19 ng L−1; Flu: 32 ng L−1) and the northern SCS (Phen: 14 ng L−1; Flu: 28 ng

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L−1),5 and significantly lower than those found in the coast of the Bohai Sea in China (Flu:

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132 ng L−1 and Phen: 449 ng L−1).48 Along the west coast of the TWS, Σ15PAHs in the

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representative estuarine (site JRDW, 145 ng L−1), harbor (site XMWH, 98 ng L−1) and

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coastal currents (site MZCC, 101 ng L−1) were significantly higher than those in the central

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TWS (62±8.3 ng L−1). Σ15PAHs values in the southern TWS (site SCSWC) were lower due

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to weaker anthropogenic inputs and dilution by the adjacent open seawater (Figures 2a, S1,

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and S2). The cluster analysis showed PAHs in the different water masses and the central

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TWS were distributed in the different groups (Figure S3), which also reflected the

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significant regional difference of the PAHs composition.

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As in the surface seawater, no significant differences were found in the total or

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individual concentrations of PAHs in the aerosol samples over the central TWS (sites

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TW01 to TW03) (Figures 2b, S1, and S2). Σ15PAHs values ranged from 23−31 and 0.2−0.4

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ng m−3 in the gas and aerosol phases, respectively. In the dominant gas phase, Phen (16±3.4

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ng m−3) displayed the highest abundance, with an average contribution of 59% to Σ15PAHs.

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Discharge from the urban agglomerations on both sides of the TWS and volatilization from

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seawater resulted in 2−5x higher Σ15PAHs values than in the ECS (8 ng m−3)14 and the

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northwestern Pacific Ocean (6 ng m−3).49 Along the west coast of the TWS, the Σ15PAHs

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value in the gas phase of the JRDW (47 ng m−3) was ~2x higher than over the central TWS

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due to significantly higher Fluo and Py inputs from the surrounding cities (Figure 2b).

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Compared with the central TWS, we found a significantly lower Phen concentration (2.3

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ng m−3) while the higher Ace, Acen and Flu concentrations in the southern TWS (Figure

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2b), which was similar to h the open SCS in our previous study.5 The different composition

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of PAHs in the central TWS was probably owing to the source differences and the effect of

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long-range transport of the air in summer (Figure S4).

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279 280

Figure 2. Relative levels of PAHs in the surface dissolved seawater and the gas phase of the air in

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the TWS.

282 283

Some studies have suggested that soil is an important natural/biogenic source of

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PAHs.50 Enhanced surface runoff and soil erosion in summer are the crucial input pathways

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of these PAHs to the coastal ocean.5, 39 In our study area, the southern TWS is influenced

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by the combined inputs of Pearl River diluted water and the SCSWC (Figure 1).51, 52

287

Therefore, the northeastward flowing SCSWC (e.g., site SCSWC) and surface runoff

288

inputs (e.g., JRDW) could be seen as two crucial sources of PAHs in the central TWS.

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These two end-members, with distinctly different concentrations and compositions of

290

PAHs, adequately mix together, driven by the confluence of the water masses and 10


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atmospheric diffusion.53 Accompanied by photo- and/or bio- degradation and the air-water

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transport of PAHs in this mixing process, PAHs displayed uniform characteristics

293

(concentrations, compositions and phase partitioning) in the central TWS.

294

According to backtrack air mass trajectories, the air of the central TWS was affected

295

by the long-range transport of air from the eastern SCS, Philippine islands, and the western

296

Pacific (Figure S4). During this transport, photo-degradation, atmospheric deposition, air-

297

water exchange of PAHs and mixing processes with terrigenous air masses resulted in

298

significant variations in PAH concentrations and compositions in the atmosphere of the

299

central TWS compared with these surrounding area (Figure 2). According to their forward

300

trajectories, PAHs in the atmosphere of the central TWS would be transported to South

301

Korea and Japan in the coming days (about 120 hours) after mixing with atmospheric PAHs

302

from the northern Chinese mainland (Figure S4). PAHs in surface seawater of the TWS

303

showed the same transport routes as PAHs in the air because of the wind-driven ocean

304

currents along the Chinese coast in summer.52

305

3.2. PAHs Transport at the Air-Water Interface.

306

The fugacity gradient of semi-volatile chemicals mathematically describes the

307

potential migration direction in which chemicals diffuse or are transported at the air-water

308

interface.9, 14, 54, 55 The fugacity ratio (fa/fw) and the fluxes of PAHs (Fa/w, ng m−2 day−1) and

309

their uncertainties were used to discuss the transport behaviors (volatilization or deposition

310

and their fluxes). The detailed description and calculations are listed in sections 5 and 6 of

311

the SI.

312

In the TWS, uncertainties of the fugacity gradient in summer were ~52% (calculated

313

in section 5 of the SI), controlled by the gas phase and dissolved concentrations of PAHs,

314

ambient temperatures and Henry’s law constant.7, 56 Therefore, 0.59 < fa/fw < 1.7 (standard

315

deviation in ln(fa/fw)=0.52) were considered to not significantly differ from phase

316

equilibrium. 3- ring PAHs showed significant volatilization across the air-water interface

317

of the TWS, which includes the dominant components of the PAH pools (Flu and Phen,

318

averaging 58% of total PAHs in the gas phase and 67% in dissolved water) (Figure 3a).

319

Fluo and Py distributions (averaging 25% of total PAHs in the gas phase and 11% in

320

dissolved water) highlighted the depositional process in the relatively open central TWS

321

and SCSWC. However, Fluo and Py distributions in the JRDW highlighted the exchange

322

equilibrium, because the local PAH emissions into the water were from the Jiulong River,

323

harbors and coastal cities. The opposite transport directions of the isomeric BaA 11



324

(volatilization) and Chry (deposition) in the TWS are a function of the differences of their

325

sources and degradation rates in air and surface water.57

326

Overall, the transfer velocities (Kol) of the individual PAHs (Table S3) in the TWS

327

averaged 0.73±0.04 m day−1, a function of wind speed and temperature.54 Determined Kol

328

values are comparable to previous reports from Chinese coasts.14 The relative uncertainties

329

of transport fluxes of individual PAHs ranged from 44% to 142%, calculated in section 6

330

of the SI, a function of Kol error propagation, compound concentrations, ambient

331

temperatures and Henry’s law constant.8, 56, 58 In the central TWS (sites TW01-TW03),

332

volatilization fluxes of the air-water exchange (Fa/w) of the dominant PAHs: Flu and Phen,

333

were as high as 14±0.9 and 12±2.7 μg m−2 day−1, respectively (Figure 3b and Table S4).

334

These fluxes are comparable to the ECS,14 lower than in the Bohai and Yellow Seas,59 but

335

higher than at the southern tip of Taiwan Island.8 This means surface seawater was the most

336

important secondary source of 3- ring PAHs to the atmosphere of the TWS. These PAHs

337

were mainly derived from riverine inputs,33 adjacent ocean inputs,5 sediment release by

338

upwelling41 and offshore oil pollution.60, 61 This further confirmed the volatile behavior of

339

LMW PAHs in the marine ocean is an important driving factor of the cross-sea transport

340

of anthropogenic PAHs.13, 14

341

In the JRDW, the volatile fluxes of LMW PAHs (Ace, Acen, and Flu, 28% of the total

342

PAHs) were 1.5−2x higher than those in the central TWS (Figure 3b and Table S4). The

343

volatile fluxes of Py, BaA, and Chry were caused by their greater concentrations in

344

seawater due to the inputs of PAHs from the Jiulong River and the surrounding cities. Fluo,

345

Py and Chry (4- ring PAHs) showed the statistical exchange equilibrium (0.59 < fa/fw < 1.7)

346

in the JRDW, but the Py and Chry still had the volatile trends with the higher fluxes (Figure

347

3a and 3b). The higher volatile flux of BaA than that in the central TWS (Figure 3b) could

348

be attributed to the higher fugacity gradient of BaA between the JRDW and the surrounding

349

urban atmosphere. In the SCSWC, the volatile fluxes of the dominant PAHs: Flu and Phen,

350

were 1.5−2x lower than in the central TWS due to the continual air-water interaction of

351

PAHs during their long-range transport from the SCS (Figures 1 and S4). In other words,

352

controlled by the surrounding input, migration and transformation characteristics of PAHs

353

in the different regions, fugacity gradient of individual PAHs in the air and water ultimately

354

determined their transport behaviors (volatilization or deposition and their fluxes) of PAHs

355

in the TWS.

356

12


Page 12 of 25

Page 13 of 25


357 358

Figure 3. Air-water interface transport of selected PAHs. (a) Fugacity ratios, the

359

horizontal gray area (standard deviation in ln(fa/fw)= 0.52) were considered to not

360

significantly differ from phase equilibrium; (b) Air-water exchange fluxes (Fa/w, μg

361

m−2 day−1) and their relative uncertainties (top x-axis, %); (c) Dry deposition fluxes

362

(Fd, μg m−2 day−1); (d) Total fluxes (Fa/w+d, μg m−2 day−1).

363 364

It has been an effort to measure the dry deposition fluxes (Fd, μg m−2 day−1) of PAHs

365

in the open ocean.62, 63 The detailed description and calculations are listed in section 7 of

366

the SI. In the TWS, the estimated dry deposition fluxes (Fd) of 3- plus 4- ring PAHs in the

367

JRDW (56 μg m−2 day−1) and SCSWC (23 μg m−2 day−1) were one order of magnitude

368

higher than those in the central TWS (1.3±0.5 μg m−2 day−1) (Figure 3c). In the JRDW, that

369

was because of the local higher particulate PAH discharge in the atmosphere; but in the

370

SCSWC, that could be due to the higher loadings of PAHs on the fine particulates with

371

larger specific surface areas in the long-range transport of SCSWC.64 In addition, because

372

of discontinuity and much shorter durations of rainfall, wet deposition is considered to be

373

at least ten times less contribution to PAHs in the air-water interface than dry deposition.65

374

Therefore, examining the combined impacts of free air-water exchange and dry deposition

375

(Fa/w+d), the transport fluxes of PAHs at the air-water interface in the central TWS showed

376

significant differences with the JRDW and SCSWC (Figure 3d). In the central TWS, the

13



377

transport of 3- and 4- ring PAHs across the air-water interface were dominated by

378

volatilization and deposition processes, respectively, based on the Fa/w values of these PAHs

379

(Figures 3c and 3d). However, Fd of Phen and An exceeded the volatile Fa/w values in the

380

SCSWC and JRDW. In the JRDW, there were higher depositional fluxes (Fa/w+d) of Phen

381

and Fluo, but other 3- and 4- ring PAHs showed primarily volatilization (Figure 3d). In the

382

SCSWC, Phen, An and 4- ring PAH fluxes were dominated by deposition at the air-water

383

interface and the LMW PAHs (Ace, Acen, and Flu) showed net volatilization (Figure 3d).

384

Similar transport characteristics of PAHs at the air-water interface have been reported in

385

many other sea areas.2, 14

386

3.3. Fossil Fuel vs. Biomass Burning Contributions of PAHs by CSRA.

387

Based on the available concentrated carbon weight (Table S5), carbon isotopic (Δ14C

388

and δ13C) measurements of four selected individual and isomeric PAHs (Flu, Phen plus An,

389

Fluo and Py) dissolved in the seawater from four water masses around the TWS were made

390

by advanced CSRA technology. The total amount of compound mass for Δ14C analysis

391

ranged from 1.8 to 5.0 μmol C, and the purities of the isolated samples were in the range

392

of 92 to 99% (Figure S5). Limited by the sample sizes of the purified PAHs, only δ13C

393

measurements of Phen plus An (−29±0.6‰) in the XMWH were obtained. The result

394

showed a similar δ13C composition as atmospheric particulate PAHs in Croatia and Greece

395

(−29 and −28‰),66 but slightly lighter than the δ13C values of atmospheric PAHs (−29 to

396

−24‰),24, 29 PAHs in central European forest soils (−25 to −23‰)26 and sedimentary PAHs

397

(−27 to −25‰).30 Although the δ13C signature has been attempted to be used to identify the

398

sources of PAHs,67-69 its application is extremely difficult because of the overlapping end-

399

member values of various PAH sources.68, 70, 71

400

In the four water masses around the TWS, the measured Δ14C contents of selected

401

PAHs ranged from −997±4‰ to −873±6‰ (Table S5 and Figure 4), which indicated

402

decayed carbon contributions, in terms of 14C content. The depleted values were explained

403

by the emission of PAHs originating from fossil contributions, including the pyrolysis of

404

fossil fuels and the release of petroleum products.3, 4 The measured Δ14C values of PAHs

405

around the TWS were comparable to those in the atmosphere of North Birmingham in the

406

southeastern US (−990 to −911‰),29 background areas of Croatia and Greece (−941 and

407

−888‰),66 oil sands in Northern Alberta, Canada (−962 to −849‰)32 and forest soils in

408

central Europe (−942 to −819‰).26 In contrast, Δ14C values of selected PAHs were more

409

depleted than those reported in sediment cores from Siskiwit (−783 to −388‰),72 14


Page 14 of 25

Page 15 of 25


410

atmospheric PAHs in Sweden (−388 and −381‰),30 residential areas of suburban Tokyo

411

(−787 to −514‰)25 and the Western Balkans (−573 to −288‰),24 which all revealed a

412

greater contribution from biomass burning. The regional differences in Δ14C values implied

413

the significant regional differences in the contribution of biomass burning and fossil fuel

414

to PAHs. To a certain extent, this also indirectly reflected the regional differences in energy

415

structure.

416

At the air-water interface of the TWS, in consideration of the net volatilization of Flu,

417

Phen and An (3- ring PAHs) and the net deposition of Fluo and Py (4- ring PAHs) (Figure

418

3d), the depleted Δ14C values of Flu, Phen and An in the water would lead to a gradual

419

increase of estimated fossil fuel contributions to 3- ring PAHs in the atmosphere. Likewise,

420

the depleted Δ14C values of Fluo and Py could also be influenced by the atmospheric

421

deposition of PAHs derived from fossil fuels or the combustion of fossil fuels.

422

423 424

Figure 4. Δ14C (‰) and ffossil fuel (%) of PAHs determined by CSRA in the water masses of the

425

TWS

426 427

Δ14C values of individual PAHs in the different water masses, in increasing order,

428

were SCSWC, MZCC, XMWH, and JRDW (Figure 4). This indicates, from lower to higher

429

values, a gradually decreasing proportion of fossil fuel contributions to the PAH pool. In

430

the SCSWC, the ocean-based anthropogenic discharge of PAHs, including potential ship

431

emissions, oil spills and produced water discharge from offshore oil platforms could

432

explain the most depleted Δ14C values. In the MZCC, the coastal southwestward current

433

could carry Δ14C depleted PAHs derived from domestic coal burning for indoor heating

434

from the coast of the ECS during winter.5 In the XMWH, the local ports and the 15



Page 16 of 25

435

surrounding cities discharged the bulk of fossil fuel derived PAHs, and the latter also

436

exported a small portion of the biomass burning derived PAHs. In the JRDW, river erosion

437

carries PAHs derived from the countryside along its path, which likely contains more PAHs

438

derived from biomass burning and causes the higher Δ14C values of PAHs than those found

439

in the other water masses. Finally, comparing the Δ14C values of each individual PAH,

440

Δ14C values of Phen plus An (3- ring PAHs) were significantly higher those of Fluo (4-

441

ring PAHs) (Figure 4). This was likely due to the mixing effects of different input pathways

442

(atmospheric deposition, particle suspension, water mass transport, etc.) which loaded

443

different proportions of biomass burning versus fossil fuel derived PAHs.

444

We estimated fractional contributions of biomass burning (i.e., fbiomass) versus fossil

445

fuel (i.e., ffossil=1−fbiomass) derived sources to individual PAHs using a two-source mixing

446

method according to the mass balance equation of

447

Δ14Cfossil (1 − fbiomass),24, 25, 29 where, Δ14CPAH is the measured

448

fraction. The radiocarbon free Δ14C value of the fossil fuel end-member (i.e., Δ14Cfossil fuel)

449

is −1000‰.29 Since there is not one specific biomass burning source to the TWS, we

450

adapted a Δ14C value of 152‰ as the modern end-member in this study.29 Therefore, the

451

calculated percent contributions of biomass burning for the individual PAHs reported here

452

were all very low in the different water masses around the TWS, with a mean of 10±0.5%

453

in JRDW, 7.5±0.7% in XMWH, 2.8±0.8% in MZCC and 1.7±0.6% in SCSWC (Table S5

454

and Figure 4). These values are very different compared with total annual global

455

atmospheric emissions, which are around 60% derived from residential/commercial

456

biomass burning.73 This discrepancy could be explained by the dominant anthropogenic

457

input sources of PAHs from the ocean or surrounding regional cities in the TWS. Primarily

458

fossil fuel derived PAHs have been found in many aquatic and atmospheric environments

459

in China.18, 74, 75 Finally, we could convert the biomass burning and fossil fuel contributions

460

to the corresponding concentrations of PAHs. That result showed the contributions of the

461

leakage and combustion of fossil fuels to Phen and An were the main factor of PAHs

462

pollution in the seawater of the Western TWS in summer (Figure 5). Detailly, in the MZCC,

463

XMWH, JRDW and SCSWC, the concentrations of fossil fuel derived Flu, Phen, An and

464

Fluo were 35, 76, 68 and 18 ng L−1, respectively; and the concentrations of biomass burning

465

derived Flu, Phen, An and Fluo were 1.2, 6.8, 8.1 and 0.4 ng L−1, respectively (Table S5).

14

C: Δ14CPAH =Δ14Cbiomass fbiomass +

466

16


14

C content of each PAH

Page 17 of 25


467 468

Figure 5. The calculated weight of fossil fuel and biomass burning derived individual PAHs in

469

the dissolved phase of seawater around the TWS.

470 471

Principal component analysis-multiple linear regression (PCA-MLR) (Figure S6) and

472

the diagnostic ratios (Figure S7) that have been used widely to identify the sources of PAHs

473

were compared in this study. For the PCA-MLR analysis in this study (see section 8 of the

474

SI), the result showed that the first principal component, with high loadings of the 3- ring

475

PAHs and Fluo and Py, contributed 92% of the PAH variability. The other principal

476

components, with high loadings of HMW PAHs, contributed to 8.0% of the observed PAHs

477

variability. However, we could not definitively assign the principal components to fossil

478

fuel or biomass burning derived sources because of the lack of a clear source spectrum of

479

PAHs. In addition, selective degradation of PAHs in the environment from their sources to

480

sinks is substantial enough to lead serious error of source identification of PAHs by

481

diagnostic ratios.19 Therefore, our work suggests CSRA is the most advantageous method

482

for identifying the sources of PAHs in the environment.

483

Next, we will further apply CSRA to the study of individual PAHs in diverse

484

environmental media (such as atmospheric particulate phase, atmospheric gas phase,

485

seawater particle phase and sediments). If that, we can investigate the migration and

486

transport behavior of PAHs in the air-water-sediment system in the coast of TWS based on 17



487

the changes of 13C and 14C compositions of PAHs.

488

ASSOCIATED CONTENT

489

Supporting Information

490

Details on sample pretreatment methods, calculations of air-water fugacity ratios, air-

491

water fluxes and their uncertainty, datasets of PAH concentrations, the backward and

492

forward trajectories of air masses, and other auxiliary figures and tables are given in the

493

supporting information. This information is available free of charge via the Internet at

494

http://pubs.acs.org.

495

AUTHOR INFORMATION

496

Corresponding Authors

497

*Xinhong Wang: Phone: +86 592 2187857; email: [email protected].

498

*Li Xu: Phone: +1 508 289 3673; email: [email protected].

499

Notes

500

The authors declare no competing financial interest.

501

ACKNOWLEDGMENTS

502

This work was supported by the National Natural Science Foundation of China (NSFC)

503

Project (41276066). We thank Meihui Lin from the Fujian Marine Forecasts for providing

504

sampling support in the Min-Zhe coastal current. We also thank the crew of the R/V Hai

505

Yang 2 and R/V Yan Ping 2 for sampling in the Taiwan Strait and Xiamen Western Harbor,

506

respectively.

507

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508 509 510 511 512 513 514 515 516

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Fossil Fuel-Derived Polycyclic Aromatic Hydrocarbons in the Taiwan

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