Polybrominated diphenyl ethers, organochlorine pesticides, and ...

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Polybrominated diphenyl ethers, organochlorine pesticides, and polycyclic aromatic hydrocarbons in water from the Jiulong. River Estuary, China: levels, ...
Environ Sci Pollut Res DOI 10.1007/s11356-015-4782-2


Polybrominated diphenyl ethers, organochlorine pesticides, and polycyclic aromatic hydrocarbons in water from the Jiulong River Estuary, China: levels, distributions, influencing factors, and risk assessment Yuling Wu 1 & Xinhong Wang 1 & Yongyu Li 1 & Miaolei Ya 1 & Hui Luo 1 & Huasheng Hong 1

Received: 26 February 2015 / Accepted: 25 May 2015 # Springer-Verlag Berlin Heidelberg 2015

Abstract Estuarine systems play an important role in the transportation and transformation of organic pollutants from rivers. Polybrominated diphenyl ether (PBDE), organochlorine pesticide (OCP), and polycyclic aromatic hydrocarbon (PAH) concentrations in water of the Jiulong River Estuary (JRE), China, were investigated to characterize their distribution, possible source, and potential ecological risk as well as the influencing factors. The total concentrations of PBDEs, OCPs, and PAHs varied from 5.2 to 12.3 pg L−1, from 29.1 to 96.4 ng L−1, and from 28.6 to 48.5 ng L−1, respectively. Their compositions were all consistent at different stations; even the input pathways were multifarious. A source analysis showed that PBDEs may come from the flame retardant usages of penta-BDE and deca-BDE; hexachlorocyclohexane isomers (HCHs) were from the use of technical HCHs, while DDTs were attributed to early residuals of industrial sources, and PAHs were mainly from pyrolytic sources. The spatial distributions of PBDEs and OCPs were quite similar with their concentrations, decreasing along the estuary and then increasing when passing the Xiamen Harbor. PAH concentrations were similar along the whole estuary, suggesting that local sources and hydrological conditions might be the influencing factors. The concentrations of these pollutants changed with tidal conditions and were positively correlated with SPM, DOC, and chlorophyll a but negatively correlated with Responsible editor: Ester Heath * Xinhong Wang [email protected] 1

State Key Laboratory of Marine Environmental Science, College of the Environment and Ecology, Xiamen University, Xiamen 361102, China

salinity. The ecological risk assessment revealed that OCPs and PAHs posed slightly higher potential risks to aquatic organism in the study area. Keywords Jiulong River Estuary . POPs . Distribution . Factors . Risk assessment

Introduction Estuaries are some of the most productive ecosystems on the Earth and critical to the life history and development of various aquatic species (Zhao et al. 2011). However, estuarine ecosystems are susceptible to anthropogenic impacts because of the hydrodynamics associated with mixing of fresh and sea water and near-shore human activities (Men et al. 2009) and are the most sensitive areas to regional response of global changes (Bianchi and Allison 2009; Wetz and Yoskowitz 2013). Various hydrophobic organic contaminants such as polybrominated diphenyl ethers (PBDEs), organochlorine pesticides (OCPs), and polycyclic aromatic hydrocarbons (PAHs) from human activities released via riverine inputs and atmospheric deposition converge in the estuaries and may result in the drastic changes of the ecological environment (Sanger et al. 1999). At the same time, the estuary also acts as a transit zone in which contaminants can be transported from rivers to deeper oceans (Men et al. 2009); a considerable amount of waste and pollutants were produced, which enter into the estuarine environment through different pathways (e.g., surface runoff, sewage and atmospheric deposition, etc.) and have extremely important influences to the estuarine and coastal environments (Luo et al. 2008). As there is deterioration of the global environment, continuous increases of pollutant inputs and the negative responses of marine

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ecosystems have become apparent; the biogeochemical behavior of organic pollutants in estuarine and coastal waters has been attracting considerable scientific interests (Paerl et al. 2006; Thompson et al. 2007; Rabalais et al. 2009). The Jiulong River is the second longest river in Fujian Province, southeast China; it has a catchment area of 14, 745 km2 and an annual average river discharge of 14,800× 106 m3 and flows into the coastal region of Xiamen and further into Taiwan Strait (TWS; Fig. 1; Zheng et al. 2011; Lin et al. 2013). The Jiulong River Estuary is a typical subtropical macro-tide estuary on the southwest coast of TWS and has been greatly impacted by human activities over the past 30 years (Yan et al. 2012). The estuary covers an area of 106 km2 and contains a water volume of 550×106 m3, is Fig. 1 Approximate location of sampling sites in the JRE

characterized by intense agricultural activities and rapid industrial development, and represents a potential major source of pollutants in Xiamen coastal waters. The rapid development of the Xiamen Economic Special Zone since 1986 has also resulted in significant stress to Xiamen Western Sea and its surrounding environments (Tian et al. 2008). Hong et al. (1995) showed that organic pollution in Xiamen waters has increased steadily in recent years and that there was significant petroleum and organochlorine contamination in Xiamen Western Sea sediments. Studies by Zhou et al. (2000), Zhang et al. (2002), Maskaoui et al. (2002, 2005), and Li et al.(2010) also confirmed the presence of these contaminants in water, pore water sediments, and organisms in the Xiamen Western Sea and Jiulong River Estuary. However, until this study, there

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was collected using a PVC water sampler, and water samples were stored in 4 L pre-cleaned brown glass jars. Each water sample was separated into the dissolved phase and particulate phase by a vermicular system with 450 °C burned glass fiber filters (GF/F, 0.8 μm, 90 mm) and then organic chemicals determined as per the method of Wu et al. (2011). Dissolved organics were extracted using a solid-phase extraction (SPE) system; the SPE cartridges (SUPELCLEAN ENVI-18) were first conditioned with methanol followed by Milli-Q water. After spiking with 2,4,5,6-tetrachloro-m-xylene, PCB209, and five deuterated PAHs (including naphthalene-d8, acenaphthene-d10, phenanthrene-d10, chrysene-d12, and perylene-d12), as internal standards for OCPs, PBDEs, and PAHs, respectively; water samples were passed through the cartridges at a flow rate of 5 mL min−1 under vacuum. Finally, the GF/F filters and SPE cartridges were stored at −4 °C. Organics trapped on SPE cartridges were eluted with 10 mL of ethyl acetate at a flow rate of 1–2 mL min L−1 using a SPE system under a vacuum pump; water was removed from the ethyl acetate extracts by pre-baked Na2SO4. Then, extracts were concentrated to be nearly dry by rotary evaporation; the solvent was exchanged into n-hexane. Samples were concentrated to about 100 μL under a stream of pure nitrogen and stored at −4 °C prior to instrumental analysis.

were little data on the levels of persistent organic pollutants (POPs) in the Jiulong River Estuary (JRE), despite this waterway being an important source of pollution to Xiamen Bay. This present study aims to assess the concentrations and spatial distribution of selected PBDEs, OCPs, and PAHs in dissolved phase from water of the JRE. Moreover, in order to demonstrate the potential factors controlling the distribution of POPs in the estuary waters, the influence of hydrodynamic conditions on the transport of PBDEs, OCPs, and PAHs and the possible regulation of environmental factors to their distributions were also investigated. In addition, possible sources of these pollutants were identified and their ecological risks were also evaluated.

Materials and methods Sampling and laboratory analysis The sampling stations along the JRE were selected in April 2008 (Fig. 1; Table 1). Based on the discharge of the Jiulong River, the wet season is from May to September, the dry season is from December to next February, and the rest of the year is considered to be the “normal” season. To study the effects of tides on pollutants in surface water and bottom water (above sediment about 0.5 m), they were fix-point sampled in low tide, middle tide, and high tide, respectively, from three state-controlled sections (north port 117.9141° E, 24.4673° N, middle port 117.8883° E, 24.4323° N, south port 117.8967° E, 24.4043° N) of the JRE. Moreover, to study the changes of pollutants with tides in the JRE, surface waters were collected in a fix-point time series for nearly one tide cycle (April 21, 2010, with high tides at 6:32 and 18:32 and low tide at 12:32) close to Jiyu Islet (JY), which is just in the mouth of the JRE and near the Xiamen Bay and was greatly impacted by the tidal front. Surface water was directly collected from the river using a stainless steel barrel, bottom water Table 1

Instrument analysis The PDBE-containing fraction was analyzed using gas chromatography (GC) coupled with mass spectrometry (GC-MS: Agilent 6890N GC-5975B MSD). The GC was equipped with a DB-5 HT capillary column (30-m×0.25-mm id, 0.1-μm film thickness, J&W Scientific), and helium was used as carrier gas at a flow rate of 1.25 mL min−1. GC operating conditions were as follows: oven heating started at 100 °C and held 2 min, ramped to 250 °C at 10 °C min−1 and then to 265 °C at 0.8 °C min−1, and increased to 325 °C at 25 °C min−1 with a final hold of 12 min. Samples (1 μL) were injected in splitless mode and the inlet was held at 250 °C. The MSD was operated

Sampling descriptions of the JRE


Water layer

Longitude (°N)

Latitude (°E)


SPM (mg L−1)

DOC (mg L−1)

Chlorophyll a (μg L−1)

J0 J1 J2 J3 J4 J5 J6 J7 J8 J9

Surface Surface Surface Surface Surface, bottom Surface Surface Surface, bottom Surface Surface, bottom

117.9187 117.9532 118.0353 118.0501 118.0515 118.0733 118.0958 118.1200 118.1356 118.1397

24.3961 24.4285 24.4404 24.4181 24.4163 24.4234 24.3826 24.3709 24.3371 24.3276

15.1 19.3 27.2 27.5 24.4 27.9 28.3 26.9 30.9 31.3

48.0 48.6 21.8 12.0 9.9 18.1 16.0 33.9 32.8 44.7

1.69 1.54 1.21 1.33 1.36 1.29 1.14 1.24 1.05 1.23

9.94 7.27 2.99 4.33 2.40 1.98 2.29 2.47 3.33 4.67

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on the chemical ion (CI) mode and the selective ion monitoring mode with ion source 280 °C. In this study, 27 PBDEs were measured: mono-BDE (1Br: BDE3), di-BDE (2Br: BDE7 and BDE15), tri-BDE (3Br: BDE17 and BDE28), tetra-BDE (4Br: BDE47, BDE49, BDE66, BDE71, and PBDE77), penta-BDE (5Br: BDE85, BDE99, BDE100, BDE119, and BDE126), hexa-BDE (6Br: BDE138, BDE153, BDE154, and BDE156), hepta-BDE (7Br: BDE183, BDE184, and BDE191), octa-BDE (8Br: BDE196 and BDE197), nona-BDE (9Br: BDE206 and BDE 207), and deca-BDE (10Br: BDE 209). The OCP-containing fraction was analyzed using GC coupled with an electron capture detector (GC-ECD: Agilent 5890 GC-ECD). The GC was equipped with an HP-5 capillary column (60-m×0.32-mm id, 0.25-μm film thickness), and high-purity nitrogen was used as carrier gas at a flow rate of 1 mL min−1. GC operating conditions were as follows: The injector and detector temperature was 250 and 320 °C, oven heating started at 90 °C and 1-min hold, ramped to 210 °C at the rate of 10 °C min−1 with 1-min hold, and then to 230 °C at 1 °C min−1 with 10-min hold, and finally, it increased to 250 °C at 1 °C min−1. Samples (1 μL) were injected in splitless mode. In this study, 25 OCPs were measured: hexachlorocyclohexane isomers (HCHs: α-HCH, β-HCH, γHCH, and δ-HCH), dichlorodiphenyltrichloroethane (DDT: o,p′-DDT and p,p′-DDT), and its metabolites (dichlorodiphenyldichloroethylene (DDE), o,p′-DDE and p,p ′-DDE, dichlorodiphenyldichloroethane (DDD), o,p′-DDD and p,p′-DDD), hexachlorobenzene (HCB), aldrin, isodrin, dieldrin, endrin, α-chlordane, γ-chlordane, heptachlor, heptachlor epoxide A/oxychlordane (Hep A/oxyCl, the peaks of heptachlor epoxide A and oxychlordane were superposed), heptachlor epoxide B (Hep B), endosulfate I, endosulfate II, methoxychlor, and mirex. The PAH-containing fraction was analyzed using GC coupled with mass spectrometry (GC-MS: Agilent 6890 GC-5973 MSD). The GC was equipped with a HP-5 MS capillary column (30-m×0.25-mm id, 0.25-μm film thickness), and helium was used as carrier gas at a flow rate of 1 mL min−1. GC operating conditions were as follows: Oven heating started at 60 °C and ramped to 300 °C at 5 °C min−1 with a final hold of 20 min. Samples (1 μL) were injected into splitless mode with the inlet held at 250 °C. The MSD was operated in the electron impact (EI) mode and the selective ion monitoring mode with ion source at 280 °C and electron voltage of 70 eV. In this study, 16 USEPA priority control PAHs were analyzed, naphthalene (Na), acenaphthylene (Ace), acenaphthene (Acen), fluorine (Flu), phenanthrene (Phen), anthracene (An), fluoranthene (Fluo), pyrene (Py), benzo(a)anthracene (BaA), chrysene (Chry), benzo(b)fluoranthene (BbF), benzo(k)fluoranthene (BkF), benzo(a)pyrene (BaP), indeno(1,2,3-cd)pyrene (IP), dibenz(a,h)anthracene (DBA), and benzo(ghi)perylene

(BghiP). Total PAH concentration was calculated as the sum of 15 PAH compounds except Na. Quality control/quality assurance In this study, all data were subjected to strict quality control procedures. 2,4,5,6-Tetrachloro-m-xylene, PCB209, and deuterated PAHs were used as surrogates and internal standards, which were spiked into the samples prior to extraction for quantifications. Surrogate standards were also used to monitor matrix effects and to compensate for losses involved in the sample extraction and workup. Qualitative ions (m/z) for PAHs are the molecular weights of each compound and for PBDEs are 79 and 81. Quantification of individual compounds was based on comparison of peak areas with those of the recovery standards using the internal standard method. The method recoveries of OCPs, PBDEs, and PAHs ranged from 62–97 %, from 74 to 126 %, and from 72 to 106 %, respectively, and the detection limits of OCPs, PBDEs, and PAHs ranged from 8 to 56 pg L−1, from 1.3 to 2.5 pg L−1, and from 41 to 214 pg L−1, respectively. Laboratory blanks, spiked blanks, and replicate samples were analyzed along with field samples. The blank experiments showed that there were no contaminations on OCPs and PBDEs, while the contaminations on PAHs were quite minor and did not affect the data analysis. Reported concentrations were corrected according to the recoveries of the surrogate standards.

Results and discussion Levels and composition patterns of POPs OCP concentrations in surface water from the JRE ranged from 39.3 to 96.4 ng L−1 (mean 60.0±17.7 ng L−1), with the highest concentration observed at station J0 and lowest at station J5 (Fig. 2). In bottom water, OCP concentrations ranged from 29.1 to 38.8 ng L−1, concentrations lower than the surface water. The individual OCPs were detected at different concentrations, with heptachlor epoxide B, methoxychlor, and mirex below detectable limits, and α-chlordane was only detected at station J9. The mix of OCPs was similar at all 10 stations of the JRE, with no obvious differences between surface and bottom water samples, perhaps due to the homogeneity of their sources or the complex hydrodynamics associated with mixing of fresh and sea water in the estuary and coastal area. HCHs and DDTs were the most abundant compounds in the JRE (Fig. 2), the concentrations of HCHs and DDTs in surface water from the JRE ranged from 11.2 to 28.2 ng L−1 (mean 18.0 ± 5.1 ng L−1) and from 13.8 to 29.7 ng L−1 (mean 22.5±5.9 ng L−1), respectively. The distributions of HCHs and DDTs were similar to OCPs (Fig. 2). These values are lower than previous studies. For instance,

Environ Sci Pollut Res Fig. 2 Concentrations of OCPs, DDTs, and HCHs in the JRE (surface water (gray column), bottom water (blank column))

Maskaoui et al. (2005) found the levels of all OCPs, HCHs, and DDTs in water from the Xiamen Western Sea and Jiulong River Estuary ranged from below the limit of detection (ND) to 2480.0 ng L −1 (mean 779.0 ng L −1 ), from ND to 352.0 ng L−1 (mean 71.1 ng L−1), and 0.2–63.2 ng L−1 (mean 14.3 ng L−1) in 1999. Zhang et al. (2002) found OCPs, HCHs, and DDTs in surface water from the JRE in 1999, in the range from 115.4 to 414.7 ng L−1 (mean 237.7 ng L−1), from 32.0 to 129.8 ng L −1 (mean 62.5 ng L −1 ), and from 19.2 to 96.6 ng L−1 (mean 48.7 ng L−1). In comparison with those reported for other estuary and coastal systems around China, the level of HCHs and DDTs in the JRE were higher than Daliao River Estuary (3.43 to 23.77 ng L−1 and 0.02 to 5.24 ng L−1) (Tan et al. 2009), Deep Bay (0.63±0.31 ng L−1 and 2.39±1.04 ng L−1) (Qiu et al. 2009), Yangtze River estuary (0.71 to 4.54 ng L−1 and 0.28 to 4.85 ng L−1) (Tang et al. 2013), and Hong Kong coastal area (0.41 and 0.94 ng L−1 and 0.77 to 5.58 ng L−1) (Wurl et al. 2006b). PBDE concentrations in surface water from the JRE ranged from 5.2 to 12.3 pg L−1 (mean 7.1±2.2 pg L−1), with the highest appearing at station J0 and lowest appearing at station

J5 (Fig. 3). In bottom water, PBDE concentrations in bottom water ranged from 5.5 to 7.2 pg L−1, concentrations similar to the surface water at the same stations. There were no obvious differences in PDBE concentrations between surface and bottom water samples. The PBDE concentration decreased initially and was then kept in steady levels. The 5Br, 6Br, and 9Br compounds dominated in the PBDE assemblages, and the compositions of PBDEs were similar at all stations (Fig. 4). The individual PBDEs were detected at different concentrations, BDE49, BDE71, BDE66, BDE119, and BDE197 were below detectable limits, and BDE77 was only detected at station J1. Oros et al. (2005) reported that congeners BDE 47, 99, and 209 were the most abundant congeners in the dissolved phase from the San Francisco Bay. However, in this study, the concentrations of BDE207 were the highest (an average 1.2 pg L−1) in all of the components, followed by BDE85, BDE3, BDE126, and BDE156 (average 0.7, 0.6, 0.6, and 0.4 pg L−1, respectively). Chen et al. (2011) found that the PBDE concentrations (including BDE28, 47, 99, 100, 153, 154, 183, and 209) in the dissolved phase from Pearl River Estuary ranged from 9.0 to 17.7 pg L−1 with a mean

Environ Sci Pollut Res Fig. 3 Concentrations of PBDEs and PAHs in the JRE (surface water (grey column), bottom water (blank column))

value of 13.6 pg L−1, which were comparable to our study. Wurl et al. (2006a) reported that the PBDE concentrations (including BDE28, 47, 99, 100, 153, 156, 183, and 209) ranged from 11.3 to 62.3 pg L−1 in the dissolved phase of Fig. 4 Compositions of OCPs, PBDEs, and PAHs in the JRE

seawater in the Hong Kong coastal area. Cetin and Odabasi (2007) reported that the mean dissolved phase PBDE concentrations (including BDE28, 47, 99, 100, 153, 154, and 209) in the summer and winter from Izmir Bay, Turkey, were 212.0

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and 87.0 pg L−1, respectively, which were much higher than our study. PAH concentrations in surface water and bottom water from the JRE ranged from 28.6 to 48.5 ng L−1 (mean 39.0± 6.2 ng L−1) and from 31.8 to 34.4 ng L−1, respectively, with the highest appearing at station J0 and lowest appearing at station J2 (Fig. 3). There was no difference in PAH profile between stations, although the PAH levels in surface water were higher than those in bottom water at the same stations. Three and four-ring PHs were most commonly observed; there were few detections of BbF, BkF, BaP, and other fivering PAHS, with IP, DBA, BghiP, and other six-ring PAHs below detectable limits (Fig. 4). Three-ring PAHs were dominant compounds with high concentrations, the percentages of Flu and Phen to total PAHs for 30 to 38 % and 26 to 33 %. Zhou et al. (2000) found that total PAH concentrations in water from the Xiamen Western Sea and Jiulong River Estuary varied from 106 to 945 ng L−1 (mean 355 ng L−1) in 1998. Maskaoui et al. (2002) found that total PAH concentrations in water from the Xiamen Western Sea and Jiulong River Estuary varied from 6960 to 26,900 ng L−1 (mean 17,000 ng L−1) in 1999. Finally, Ya et al. (2014) reported the dissolved PAH concentrations in subsurface water along the coast of Xiamen ranged from 62.0 to 341.9 ng L−1 (mean 182.4 ng L−1) in 2005. Compared with several years ago, PAH concentrations in the JRE have dramatically decreased. In comparison with those reported for other estuaries around China, the level of PAHs in the Jiulong River Estuary was lower than that in the Daliao River Estuary (139.2–1717.9 ng L−1) (Men et al. 2009) and Yellow River Estuary (469.8–1187.6 ng L−1) (Lang et al. 2008) but similar to that in the Pearl River estuary (12.9– 182.4 ng L−1) (Luo et al. 2008). Factors affecting the spatial variability of POPs The levels of OCPs, PBDEs, and PAHs in the estuary progressively decreased as riverine inputs into the JRE diluted their concentrations until Jiyu Islet, near the mouth of the JRE, where concentrations increased as a result of new inputs from the Western Xiamen Harbor, including urban storm water runoff, sewage discharges, and port activities. Most POPs are hydrophobic and effectively insoluble and prefer to associate with colloids, dissolved organic matter, or suspended particle matters (SPMs) in water (Shi et al. 2007). Characteristics of water that can influence the behavior and fate of POPs in the water column include dissolved organic carbon (DOC), SPM content, and salinity (Guo et al. 2009). Sorption to colloids and DOM could increase the total apparent dissolved concentration of POPs, which includes freely dissolved POPs, DOCassociated POPs, and POPs associated with colloidal or finegrained particles (Fernandes et al. 1997). The correlations between the OCPs; PBDE concentrations; and SPM, DOC, chlorophyll a, and salinity were all significant (Table 2), although

Table 2 factors

Pearson’s correlation matrix for the POP concentrations and SPM


Chlorophyll a

Salinity r=−0.6808**









r=0.6611* p=0.037

r=0.6670* p=0.012,

r=0.8919** p=0.001

r=−0.8919** p=0.001











r is the correlation coefficient, and p is the significance level *Correlation is significant at the 0.05 level (two-tailed); **correlation is significant at the 0.01 level (two-tailed)

the correlations between the PAH concentrations and environmental factors were not. SPM in water is from terrestrial runoff, atmospheric precipitation, sediment resuspension, and biogenic particulate matters from the excreta or remains of bacteria and plankton. In the normal season, there were positive correlations between SPM and OCP, PBDE, and PAH concentrations in the JRE (Table 2, Fig. 5), especially OCPs and PBDEs. As the concentration of SPM increased, the concentrations of dissolved and particle phase OCPs and PBDEs also increased. The reason is that the higher the SPM concentration in the estuary, the higher the level of colloidal organic matter and thus the higher the concentration of apparently dissolved POPs (Schrap et al. 1995). DOC is one of the most important factors influencing the behavior, fate, and toxicity of dissolved phase POPs in water (Haitzer et al. 1998; Tremblay et al. 2005). Gale et al. (1992) found that DOC easily enriched POPs through adsorption and complexation due to the low water solubility and high-lipidsolubility characteristics of POPs. Similar to SPM, positive correlations were found between DOC content and OCPs and PBDEs in water of the JRE (Table 2, Fig. 5), i.e., as DOC content increased, the content of the dissolved contaminant in water also increased. So, for this kind of hydrophobic organic pollutants, the dissolved organic matter content is one of the most important factors that influence their distributions in the JRE. Chlorophyll a concentration is used as a surrogate for algal biomass. In that context, positive concentrations were found between chlorophyll a contents and OCPs, PBDEs, and PAHs in water of the JRE (Table 2, Fig. 5). Li et al. (2011) found that the photosynthetic rate based on chlorophyll a at noon time in the JRE increased from the low-salinity zone to turbidity front and then decreased in the mixohaline zone. Thus it can be seen that chlorophyll a also influences the distribution of organic pollutant in surface water of the JRE. As salinity increases, the solubility of many chemicals decreases. This “salting-out effect” causes organic pollutants to

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Fig. 5 Correlations between POPs and environmental factors in the JRE

be adsorbed to suspended particles and then deposited onto the sediment (Pearson et al. 1996). Negative correlations were found between salinity and OCPs and PBDEs in water of the JRE. One of the reasons for this effect was the salting out in the estuary; another was the diluting effect of seawater. This is consistent with Luo et al. (2008) who found that salinity correlated to the particle-water portioning of PAHs in the Pearl River estuary and its coastal water. Ou et al. (2010) found that the “salting-out effect” is one of the important factors that influence the seasonal variations of dissolved PAHs in the Yangtze River estuary. So, salinity also influences the distribution of organic pollutant in surface water of the JRE. The salinity profiles of OCPs, PBDEs, and PAHs showed decreasing concentrations with increasing salinity, suggesting possible runoffs from agricultural land along the Jiulong River

watershed. While the concentrations of dissolved PAHs remained relatively constant, suggesting that not only this contaminant was from river runoff or seawater outfalls but also it was mainly derived from land-based atmospheric deposition (Wu et al. 2011). Tides and currents act a very important role on water transport and pollutant migrations in estuaries and gulfs. Tide is a major factor affecting the mixing of freshwater and seawater, along with wind and advective exchange. The JRE has a regular semidiurnal tide, with reciprocating flow and salinity in the range 3.4 to 31.8. The ebb and flow of tides make for a reciprocating motion of pollutants carried by water in the estuary that can accelerate the diffusion of pollutants, and the dynamic action in the tidal zone produces laminar flow, which has a significant influence on the transport of pollutants. In

Environ Sci Pollut Res Fig. 6 OCPs and PAHs in surface waters from the JRE in different tide periods

this study, OCP and PAH concentrations were high at low tide but low at high tide (Fig. 6), which indicates that the pollutants in water from the JRE were mainly influenced by the river runoff at low tide, while the strong currents from the open sea have a dilution effect at high tide. However, the PAH concentrations in surface water from the north port were high at both high and low tides, indicating that the PAHs originated from the rivers at low tide but Xiamen harbor at high tide. The variations of dissolved OCP and PAH concentrations in surface water near the Jiyu Islet were fluctuant and had similar trends over time (Fig. 7); it was found that the high OCP and PAH concentrations appeared both in the highest tide period and the lowest tide period. The JY water was mainly influenced by the discharges of pollution sources from Xiamen Harbor in the high-tide period because the industrial and shipping activities were frequent, which could introduce amounts of organic pollutants to Xiamen Bay, while the JY water in the low-tide period was from riverine runoff with lower pollutant concentrations compared to the Xiamen Harbor. Furthermore, the low OCP and PAH concentration during the falling and rising periods may attribute to the amount of clean seawater from the open sea. Therefore, the OCP and PAH concentration in surface water from the mouth of the JRE was affected by the dual actions of riverine runoff from the Jiulong River watershed and the strong tide from the open sea. Sources of organic contaminants Compositional variations of HCH isomers in the environment generally indicate varied sources of contaminants (Doong et al. 2002). Usually, technical HCH contains 60–70 % of αHCH, 5–12 % of β-HCH, 10–12 % of γ-HCH, and 6–10 % of δ-HCH, respectively (Willett et al. 1998; Breivik et al. 1999), while lindane contains 99 % of γ-HCH. Among the isomers of HCHs, β-HCH is the most stable and relatively resistant to microbial degradation (Willett et al. 1998). The predominance of β-HCH can be used for exploring the source of historical

usage of HCHs (Law et al. 2001). In water, the solubility of the various HCHs is δ > γ > α > β. In the JRE, the α-, β-, γ-, and δ-HCH isomers contributed about 15–35 % (mean 26 %), 0–33 % (mean 16 %), 5–10 % (mean 7 %), and 40–68 % (mean 51 %), respectively. The lack of β-HCH perhaps suggests continuous use of HCH in the surrounding catchment. Ratios of α/γ-HCH between 3 and 7 represent the technical mixture, while those close to 1 indicate the use of lindane (Li et al. 1998). In the JRE, α- to γ-isomer ratios (α/γ ratios) of dissolved phase samples were from 1.9 to 5.4 (Table 3), which indicates that atmospheric deposition may be another possible source of HCH residues in this area (Walker et al. 1999). Generally, technical DDT contains 80– 85 % p,p′-DDT and 15–20 % o,p′-DDT (Metcalf 1973). DDTs can be biodegraded into DDE under aerobic condition and to DDD under anaerobic environment (Hitch and Day 1992), so the relative concentration of the parent DDT compared to its biological metabolites (DDD and DDE) can be used as indicative indices for assessing the possible pollution sources. Specifically, the ratios of DDT/(DDE + DDD) and DDD/DDE can be used as indicative indices for assessing the long-term weathering and biotransformation of DDT under various redox conditions, respectively (Doong et al. 2002). At all stations, o,p′-DDT concentrations were less than p,p′-DDT, suggesting that DDTs were mainly from industrial sources in the environment of the JRE. The DDD/DDE and DDT/(DDE + DDD) ratios (Table 3) suggest that DDT has not been used recently in the region and that the detected DDT concentrations are derived from aged and weathered agricultural soils or other residuals. The constituents of three commercial PBDE mixtures are designated as penta-BDE, octa-BDE, and deca-BDE. BDE209 is the main component in deca-BDE industrial products (>90 %) (Rayne and Ikonomou 2002), and its presence reflects the use of deca-BDE as the major brominated flame retardant. However, recent studies found that the photosensitive BDE209 is easily degraded (Ahn et al. 2006) and tends to

Environ Sci Pollut Res Fig. 7 Correlations between OCPs, PAHs, and tidal current in the JRE

industrial discharges, petroleum spills, fossil fuel combustion, and automobile exhaust, as well as via non-point sources, such as urban runoff and atmospheric fallout (Countway et al. 2003; Shi et al. 2005). Anthropogenic sources of PAHs are classified as pyrolytic PAHs (incomplete combustion of organic matters, anthropogenic industrial activities, or natural fire) and petrogenic PAHs (crude oil or unburned petroleum and its refined product) (Zhang et al. 2005) and may be identified by ratios of individual PAH compounds based on peculiarities in PAH composition and distribution pattern as a function of the emission source (Zhang et al. 2003; Guo et al. 2007). Isomeric ratios of PAHs are classical indices and have been widely used to evaluate the possible source categories of PAHs in the environment. Several PAH congener ratios have been selected as indicators that exhibit the best potential to elucidate/distinguish natural and anthropogenic sources or pyrolytic and petrogenic anthropogenic sources. According to many references (Yunker et al. 2002; Tobiszewski and Namieśnik 2012), the concentration ratios of some selected PAHs are considered to be characteristics of their sources. In this study, PAH isomer pair ratios An/(Phen + An), Fluo/(Fluo

degrade in anaerobic condition (Gerecke et al. 2005), with the nona-BDE (BDE-206, BDE-207, and BDE-208) being the dominant degradation product. Therefore, the relative abundance of nona-BDE in water from the JRE also indicated the use of deca-BDE commercial products in this region, which suggested that deca-BDE formulation was the predominant commercial PBDE product used in the JRE. And, it agreed with the fact that technical deca-BDE mixtures were the dominant PBDE formulation used in China (Mai et al. 2005; Zou et al. 2007). BDE47, BDE99, BDE100, BDE153, and BDE154 are the major components of penta-BDE commercial products (Su et al. 2007). In the JRE, PBDE 99, PBDE 85, and PBDE 126 are also the dominant compounds, inferring that they were from penta-BDE commercial products. The relative high abundance of BDE156 and BDE184 was deemed to be octa-BDE products (Song et al. 2004). PAHs have natural sources (oil seeps, forest and prairie fires, volcanic eruptions and diagenesis) and anthropogenic sources (fossil fuel burning, biomass combustion, and industrial processes) (Tobiszewski and Namieśnik 2012). PAHs are introduced into the environment mainly via municipal/ Table 3 Source identifications of POPs in the JRE












α-HCH/γ-HCH DDT/(DDE + DDD) DDD/DDE o,p′-DDTs/p,p′-DDTs An/(Phen + An) Fluo/(Fluo + Py) BaA/(BaA + Chry)

1.90 0.11 1.15 0.63 0.13 0.44 0.25

5.37 0.34 1.08 0.45 0.19 0.48 0.26

4.37 0.51 0.63 0.11 0.15 0.56 0.35

4.62 0.42 0.90 0.29 0.11 0.57 0.27

3.52 0.18 0.58 0.42 0.15 0.57 0.23

2.54 0.39 0.92 0.33 0.12 0.58 0.31

2.88 0.33 0.91 0.46 0.14 0.58 0.31

4.23 0.36 0.82 0.51 0.15 0.55 0.30

3.72 0.63 1.01 0.36 0.11 0.59 0.30

4.12 0.46 0.99 0.50 0.11 0.61 0.27

Environ Sci Pollut Res

+ Py), and BaA/(BaA + Chry) were used to assess different sources. Generally, the An/(Phen + An) ratio of 0.1 is used to differentiate between pyrolytic and petrogenic origins, and the Fluo/(Fluo + Py) ratio of 0.5 and the BaA/(BaA + Chry) ratio of 0.35 have been proven as the transition points of these two sources. Specifically, the Fluo/(Fluo + Py) ratio 0.5 hints grass, wood, or coal combustion. Similarly, the BaA/(BaA + Chry) ratio 0.5 infers grass, wood, and coal combustion (Yunker et al. 2002; Zhang et al. 2005). As shown in Table 3, in the JRE, An/(Phen + An) ratio was 0.11 to 0.19 (mean 0.13), Fluo/(Fluo + Py) ratio was 0.44 to 0.61 (mean 0.56), and the BaA/(BaA + Chry) was 0.22 to 0.35 (mean 0.29), suggesting that PAHs mainly originated from the mixed inputs of pyrogenic and pyrolytic sources (combustion), such as the diesel direct input or incomplete combustion of oil, coal and wood, which mainly represents the land-based organic matter via river runoff and waste discharge. Moreover, the values of PAH isomer ratios in the surface water from the JRE were very close to the values in PM10 from aerosol samples from the JRE (Wu et al. 2012), indicating that the atmospheric deposition also contributed to dissolved PAH input sources in water of the JRE. Ecotoxicological risk assessment In this study, the risk evaluation of water environment follows the water quality standards developed by environmental protection ministries. Results from this study show that the concentrations of OCPs in the JRE water samples were below the environment quality standard for surface water (GB38382002) established by the State Environment Protection Agency (SEPA) of China (

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