Occurrence and potential health risks assessment of polycyclic aromatic hydrocarbons (PAHs) in different tissues of bivalves from Hainan Island, China

Food and Chemical Toxicology 136 (2020) 111108

Contents lists available at ScienceDirect

Food and Chemical Toxicology journal homepage: www.elsevier.com/locate/foodchemtox

Occurrence and potential health risks assessment of polycyclic aromatic hydrocarbons (PAHs) in different tissues of bivalves from Hainan Island, China

T

Haihua Wanga,b, Wei Huangb, Ying Gonga,b, Chienmin Chenc, Tengyun Zhanga,b, Xiaoping Diaoa,b,d,∗ a

College of Ecology and Environment, Hainan University, Haikou, 570228, China State Key Laboratory of Marine Resource Utilization in South China Sea, Hainan University, Haikou, 570228, China Department of Environmental Resources and Management, Chia Nan University of Science and Pharmacy, Tainan, 717, Taiwan, ROC d School of Biology, Hainan Normal University, Haikou, 571158, China b c

ARTICLE INFO

ABSTRACT

Keywords: PAHs Bivalves Bioaccumulation Daily intake Health risks

The levels of 16 PAHs were determined in the adductor, gills, gonads, hepatopancreas and mantles of the pearl oyster (Pinctada martensii) and the mussel (Perna viridis) collected from coasts of Li'an and Xincun Bays. The levels of ΣPAHs ranged from 597.1 to 2332 ng g−1 d w in the various tissues of bivalves. The pyrolytic source played an important role in the local coastal environment. Significantly higher levels of M-PAHs and H-PAHs were observed in Pinctada martensii than in Perna viridis. The ΣPAHs at different tissues showed the following order from high to low: mantles > hepatopancreas > gonads > gills > adductor. When levels of individual PAHs in the five bivalve tissues have been compared with each other, high correlations have been found (r2 = 0.793–0.975). A general trend was observed that log transformed BSAFs declined with increase of Kow values. The estimated amount of ΣPAHs via ingestion of oyster and mussel varied from 1.35 × 10−21.70 × 10−1 and 2.15 × 10−2-1.91 × 10−1 μg kg−1 body weight day−1, respectively. The THQs and CRs calculated for regular consumption of raw bivalves were in the acceptable ranges and may not pose health risk concerns. But for certain population with higher consumption rate for PAHs contaminated bivalves, cautions should be taken for their higher cancer risk.

1. Introduction PAHs are one group of organic contaminants containing two or more fused benzene rings. Some PAHs show very high mutagenic, carcinogenic activity and toxicity (Pan et al., 2017; White, 2002). PAHs are produced from a variety of anthropogenic and natural sources. Anthropogenic sources of PAHs include coal tar, burning of fossil fuel, wood, garbage, used lubricating oil and oil filters (Kaushiket al., 2006). Natural sources are forest and rangeland fires, volcanic eruptions, oil seeps and exudates from trees (Kaushiket et al., 2006). PAHs are very prevalent and persistent in the environment, usually present in complex mixtures (Kim et al., 2013; Montuori et al., 2016; Sun et al., 2018). Due to lipophilic characteristics and resistance to biodegradation of these contaminants, they tend to accumulate in organisms through bioconcentration and biomagnification, and finally resulting in health risk for humans through dietary intake (Oliveira et al., 2018; Rengarajan et al., 2015).

∗

Seafood can be reflected the local marine environmental contamination level (Abdel-Shafy et al., 2016). Global seafood, mainly comprised of fish and shellfish, is consumed by human worldwide, and the consumption rate has more than doubled in the past 50 years, to over 20 kg per capita per year in 2014 (FAO, 2016). In China, the rate has been reported to be doubled (OECD-FAO, 2019). Seafood also provided more than 3 billion people with at least 20% of their intake of animal protein (FAO, 2016). In China, national shellfish aquaculture production in 2006 reached 11.136 million tonnes, and accounted for about 77% of mariculture production for that year (Wei et al., 2010). Consumption of bivalve mollusks in China, especially in coastal areas, has increased in recent years in response to the higher availability from either wild catch or aquaculture. Molluscs are the major marine species farmed and account for approximately 78.60 percent in China (FAO, 2019). As bottom feeders, bivalves frequently come in contact with water

Corresponding author. College of Ecology and Environment, Hainan University, Haikou, 570228, China. E-mail address: [email protected] (X. Diao).

https://doi.org/10.1016/j.fct.2019.111108 Received 14 August 2019; Received in revised form 27 December 2019; Accepted 29 December 2019 Available online 31 December 2019 0278-6915/ © 2019 Elsevier Ltd. All rights reserved.

Food and Chemical Toxicology 136 (2020) 111108

H. Wang, et al.

and sediment containing hydrophobic contaminants. Not only with economic values, bivalve populations but also play a significant ecological role in estuarine and coastal habitats. Variations in population dynamics and reproductive output of bivalves are highly sensitive to environmental disturbances (Weinberg et al., 1997). So that benthos bivalve molluscs have been used as biological indicator organisms to monitor marine environmental contaminants (Richardson et al., 2008; Soriano et al., 2006; Wang et al., 2011). In general, for some hydrophobic organic pollutants such as PAHs, higher trophic organisms exhibit relatively higher chemical concentrations than the lower ones (biomagnification), and the condition is primarily influenced by the initial pollutant levels occurring in the ambient environment (e.g. water and sediment) or foods (e.g. phytoplankton) (Deeb et al., 2007; Mashroofeh et al., 2015). PAHs pollution in bivalves has become a significant worldwide concern, not only because of the threat to bivalves, but also due to the health risks to humans associated with their consumption. Many studies found various PAHs concentrations in bivalves (Deeb et al., 2007; Mashroofeh et al., 2015). Some researchers have highlighted the importance of biological factors influencing the organic pollutant dynamics in bivalves, for example, species, habitat, feeding strategy, lipid content, trophic level and seasonality (Vieweg et al., 2012; Wang et al., 2008, 2017). There are also other studies focused on exploring the toxicological effects of PAHs in different tissues of bivalves (Chen et al., 2016; Song et al., 2017; Xie et al., 2017). Furthermore, bivalves accumulated pollutants such as PAHs from the aquatic environment representing a potential health risk for consumers. Different levels of PAHs in bivalves (Arca senilis) from southern Nigeri were detected by Moslen's group and for total PAHs, its human health risks (incremental life cancer risk) due to consumption of the species was found to exceed 10−5, a level considered to be of concern (Moslen et al., 2019). A similar magnitude of cancer risk (10−5) was reported in two case of ingestion of PAHs contaminated oysters (Crassostrea tulipa) collected in three Ghanaian coastal ecosystems (Kofi et al., 2005), and clams (Donax trunculus) caught in the Catania Gulf, Italy (Ferrante et al., 2018). Yu et al. (2016) investigated the levels, spatial and temporal trends of 16 polycyclic aromatic hydrocarbons (PAHs) in oyster Crassostrea rivularis along the Hainan Island coasts, and found various concentrations of PAHs in oysters with an average of 856.7 ng g−1 dw. In the same study, the investigators also reported that the cancer risks from consumption of PAHs contaminated oysters from Basuo Habor exceeded 10−5 (Yu et al., 2016). It should be noted that Hainan Island, with very limited industrial activities, is considered a relatively pristine area in China. Located in southeastern coast of Hainan Island, Li'an and Xincun

Bays not only have two fishing ports which have a large number of marine aquaculture, but also have heavy residential and marine recreational activities. Xincun Bay also has one of the biggest seafood markets on the Hainan Island. The two bay areas are enclosed seas each with one narrow entrance, limited water exchange from the open water into the bays and human activities with heavy traffics in that area results in concerns of marine pollution and human health risks associated with consumption of contaminated seafood. Currently, there is no data related to the levels, intake or potential health risks of PAHs due to consumption of seafood caught in this area. Besides, based on the statistics, not only did Chinese consume more fish and shellfish than the people of other countries (FAO, 2019), but coastal residents also have access to and preference for marine fish and shellfish products more. These will lead to a higher risk by contaminants in seafood. The overall aims for the present study were: (1) to investigate the levels and possible sources of 16 PAHs in different tissues of two species of bivalves mostly found in the Li'an and Xincun Bays water, namely pearl oysters (Pinctada martensii) and mussels (Perna viridis); (2) to evaluate the daily intake of PAHs through consumption of the two bivalves and its corresponding health risks. The information obtained will help us to better understand the environmental transport and fate of PAHs in the area studied and to serve as basis for sound and integrated environmental management in the future. 2. Materials and methods 2.1. Study area Both the Li'an and Xincun Bay areas have important fisheries on Hainan Island, and are also the first designated seagrass protected areas in China (Tu et al., 2016). With a total area of 23.2 km2, the area locates in the southeastern of Hainan Island (between 18°23’ to 18°26’ N and 109°58’ to 110°03’ E) (Fig. 1 and Table S1). 2.2. Collection of samples All samples were collected manually from the Li'an and Xincun Bays during September 2016. Three seawater and three sediment samples from each three points were collected at each the three sites located, from the water's edge to the center. Seawater samples were stored in brown glass bottles and were transported to the laboratory in accordance with the Offshore Marine Areas Monitoring Technical Specifications (HJ422-2008). A stainless steel sample grabber was used to collect surface sediment samples (0–10 cm), which were stored in

Fig. 1. Map of the sampling region and the Li'an and Xincun Bays located in the southeastern coast of Hainan Island. 2

Food and Chemical Toxicology 136 (2020) 111108

H. Wang, et al.

Oral reference dose (available only for Ace: 6.0 × 10−2, Flu: 4.0 × 10−2, Flua: 4.0 × 10−2, Ant: 3.0 × 10−1, Pyr: 3.0 × 10−2, Nap: 2.0 × 10−2, mg kg−1 day−1) (IRIS); AT is averaging time for exposure (EF × ED); CSF is the oral cancer slope factor (BaP: 7.3 mg kg−1 day−1), only BaA, Chy, BbF, BkF, BaP, InP, DahA for carcinogenic (USEPA, 1993), was estimated using the BaP CSF after converting to their toxic equivalency factors (TEFs) (Table S4); An example of the Calculation of THQ and CR for the consumption of bivalve tissues is showed in Tables S5 and S6, respectively.

aluminum containers. The seawater and sediment samples were refrigerated during transport to the laboratory and kept at 4 °C and −20 °C, respectively, before analysis. Similar sizes of pearl oysters (Pinctada martensii) and mussels (Perna viridis) were bought from fishermen upon collection from the two bays, and then immediately taken to the laboratory in freezer bags. Individual specimen was sacrificed with dissecting sets. Samples of mantles, gills, gonads, hepatopancreas, adductor muscle and total soft tissues were quickly removed and washed with distilled water. The tissues of specimens from each bay were pooled as composite samples and immediately refrigerated (−20 °C) until being freeze-dried and homogenized for further extractions.

2.6. Data analysis Excel and Origin (version 8.0) were used for data statistical analysis. One-way analysis of variance (ANOVA) with Duncan's method was conducted to determine significant differences among the tissues' PAH concentrations in bivalves. Independent samples t-test was conducted to determine the significant difference between two bivalve species analyzed. The correlation between the data was analyzed by the linear fit. A 5% significant level and 1% extremely significant level were used for the analysis. In this study, the PAHs concentrations of water and sediment in Li'an and Xincun Bays were pooled respectively to represent the PAHs concentration of surrounding environment where bivalves lived. Comparisons among fingerprints were shown by grouping PAHs in three classifications: high molecular weight PAHs (H-PAHs) including > 4-ring PAHs [Benzo[b]fluoranthene (BbF), Benzo[k]fluoranthene (BkF), Benzo[a]pyrene (BaP), Indeno[1,2,3-cd]pyrene (InP), Dibenz[ah]anthracene (DahA), Benzo[ghi]perylene (BghiP)], median molecular weight (M-PAHs) including 4 ring PAHs [Fluoranthene (Flua), Pyrene (Pyr), Benz[a]anthracene (BaA), Chrysene (Chy)] and low molecular weight PAHs (L-PAHs) including ≤ 3-ring PAHs [Naphthalene (Nap), Acenaphthylene (Acy), Acenaphthene (Ace), Fluorene (Flu), Phenanthrene (Phe), Anthracene (Ant)].

2.3. Samples extraction and cleanup Seawater samples were extracted using the SPE method (Li et al., 2015a). Determination of PAHs in sediment was performed based on the matrix solid-phase dispersion method with slight modifications (Cui et al., 2015). Extraction of PAHs in the bivalves was analyzed using the same method of (Cui et al., 2015), with some modifications. A more detailed description of the extraction and cleanup procedure is given in Section S1 of the Supplementary Material, including the chemicals and reagents and quality control of the methodology used for the quantification of PAHs in all samples. 2.4. BSAF calculation BSAFs in bivalves were calculated in this study in order to understand the bioaccumulation of PAHs in bivalves. The BSAF is defined as the ratio of the individual PAH (or total) lipid-normalized concentration in the bivalve tissue and its organic carbon normalized concentration in the sediment. The BSAF can be calculated using the following equation (Thorsen et al., 2004):

3. Results and discussions

BSAF = (Cb/ fl )/(Cs / foc )

The levels of the 16 PAHs determined in the all samples collected are summarized in Tables 1 and S7, while their average concentrations in different bivalve tissues are shown in Fig. S1, respectively.

Where Cb is the individual PAH concentration in bivalve tissue (ng g−1 d w), fl is the bivalves fraction (g g−1 d w), Cs is the individual PAH concentration in sediment (ng g−1 d w), and foc is the sediment fraction of organic carbon (g g−1 d w).

3.1. PAHs in water and sediment

2.5. Risk assessment

The average concentrations of ΣPAHs in Li'an and Xincun Bays' seawater were 76.26 ng L−1 and 522.9 ng L−1, respectively. The composition patterns of PAHs in the seawater (Fig. 2) indicates that 2-, 3- and 4-ring PAHs were abundant in the two sampling areas, representing over 98% of all PAHs in average. In contrast, 5- and 6-ring PAHs were observed in lower levels, accounting for only 1.5% of ΣPAHs. These findings were in accordance with previous studies (Chen et al., 2007; Li et al., 2015a). The average concentrations of ΣPAHs in the Li'an and Xincun Bay sediments are 929.2 ng g−1 and 899.8 ng g−1, respectively. In terms of the compositional profile of different PAHs in sediments at the two sampling bays, 3- and 4-ring PAHs were dominant, accounting for 61% and 28% of ΣPAHs, respectively (Fig. 2). Phe, Flua and Pyr were the dominant PAHs, which is partially in accordance with the results obtained in studies carried out in other coastal areas (Ko et al., 2014; Tam et al., 2001). The ΣPAHs levels detected in seawater in this study were not high compared to those found in other coasts in Hainan, such as the Yangpu Bay and Haikou Bay (P. Li et al., 2015; Y. Li et al., 2015), and were close to the results from other studies in other areas such as Western Taiwan Strait and Saronikos Gulf (Valavanidis et al., 2008). It is presumed that acute toxicity may not be occurred to certain organisms exposed to the levels of ΣPAHs (< 10 μg L−1) found at the two locations (Table S7) (Law et al., 1997). It is noted that the ΣPAHs concentration in the Xincun’ sample (522.9 ng L−1) was dramatically higher (p < 0.01) than that from Li'an's (76.26 ng L−1). It is probably because Li'an Bay is

For a better risk assessment due to consumption of bivalves, we applied two average consumption rates for adults based on the world (2.50 kg per capita per year) or Chinese averages (8.85 kg per capita per year) for the year of 2013 (FAO, 2019) in this study. Exposure daily intake (EDI), the target hazard quotient (THQ) and carcinogenic risks (CR) were calculated to determine non-carcinogenic effects and carcinogenic effects. The equations are the following (Oliveira et al., 2018):

Cb×IR BW

EDI = THQ =

CR =

EF× ED × IR × Cb RfD× BW×AT

EF×ED × IR × Cb×CSF BW × AT

Where Cb is the concentrations of PAHs in bivalve (μg kg−1 w w) (Table S2); an example of the conversion of the concentrations of PAHs in bivalve dry weight into the concentrations of PAHs in bivalve wet weight according to water content is presented in Table S3; IR is the average daily ingestion rate of molluscs (world: 6.85 × 10−3 and China: 24.25 × 10−3 kg day−1), provided by FAOSTAT on the “Food Supply-Livestock and Fish Primary Equivalent” database (FAO, 2019); BW is the body weight of adults (70 kg); EF is the exposure frequencies of 365 days year−1; ED is the exposure duration of 70 years; RfD is the 3

Food and Chemical Toxicology 136 (2020) 111108

H. Wang, et al.

Fig. 2. Composition pattern of PAHs in seawater, in sediment and in the bivalves from the Li'an and Xincun Bays.

mainly a fishery area, and PAHs are primarily contributed by local fishing and some residential activities. In contrast, Xincun Bay area is not only a fishing ground, but also has one of the biggest seafood markets on Hainan Island, with various and much higher human activities including road and water traffics. PAH levels may be due to accumulation of various sources of input such as petroleum combustion, vehicle exhaust emissions, domestic and commercial human activities in this region. Unexpectedly, there was no significant difference (p > 0.05) between ΣPAHs levels in Li'an and Xincun Bays' sediment samples. The similarity in ΣPAHs levels found in sediment in the two areas can probably be explained by total organic carbon (TOC). TOC is a crucial factor influencing the levels of organic pollutants in the sediment (Liang et al., 2007; Ranjbar et al., 2017; Sun et al., 2018). TOC content (6.16%) in the sediment samples from Li'an Bay was higher than that from Xincun Bay (1.55%) (Table S1), which consequently lead to more PAHs retained by the sediment even with less input of the contaminants.

mantles, respectively, while for L-PAHs, the same trends were observed as those for ΣPAHs. The levels of H-PAHs in tissues of Perna viridis were in the following order from high to low: hepatopancreas > gonads > adductor > gills > mantles; however, the levels of ΣPAHs, L-PAHs and M-PAHs in tissues of Perna viridis followed the same trends as those observed in Pinctada martensii. All sixteen PAHs analyzed were detected in all the tissues of both bivalves. The discrepancies between the mean concentrations of ΣPAHs and of 2-, 3-rings PAHs (L-PAHs) found in two species of bivalves were not statistically significant. However, higher levels for 4-rings (M-PAHs) were found in Pinctada martensii than in Perna viridis (p < 0.05), as well as for 5-, 6-rings PAHs (H-PAHs) (p < 0.01), which may be due to the difference in feeding habit between the two bivalves. Although most of the bivalves are generally benthic feeders, food composition would be different between species. For example, the pearl oyster feeds mainly on bacteria, algal picoplankton and nanoflagellates (Loret et al., 2000; Tomaru et al., 2000), while the preferred food for mussel are copepod, crab, amphipod and shrimp larvae, algae and phytoplankton, as well as detritus and some miscellaneous items (Ashraful et al., 2009). The uptake rate of individual PAH by ingestion depends on level of the PAHs in food, structure of guts, feeding rate, and duration of exposure (Logan, 2007). With relatively higher values of the octanol/water partition coefficient (log Kow > 5), PAHs will enter bivalves via direct contact of the organisms with water, sediment or by ingestion of food (Logan, 2007; Vives et al., 2004). A previous study has shown that the food chain was one of potential factors that affect bioconcentration of PAHs in fish (Wang et al., 2012).

3.2. PAHs in bivalve tissues Table 1 show that the average concentrations of ΣPAHs in different tissues ranged from 597.1 to 2332 ng g−1 for pearl oysters and from 818.5 to 2153 ng g−1 for mussels. The levels of ΣPAHs, M-PAHs and HPAHs in various tissues of Pinctada martensii showed the following, ordered from high to low: mantles > hepatopancreas > gonads > gills > adductor, mantles > gills > hepatopancreas > gonads > adductor and hepatopancreas > gonads > gills > adductor > 4

Food and Chemical Toxicology 136 (2020) 111108

322.3 ± 108.3 43.98 ± 4.54 42.55 ± 2.84 158.8 ± 16.45 1174 ± 40.17 128.7 ± 13.06 147.6 ± 15.31 183.7 ± 24.09 28.92 ± 6.24 58.84 ± 19.70 17.27 ± 3.30 6.83 ± 1.37 4.11 ± 0.19 7.36 ± 0.39 5.62 ± 0.23 1.14 ± 0.20 1871 419.1 42.32 2332 3.54 ± 0.74 78.41 ± 1.24

90.58 ± 23.20 24.41 ± 2.43 23.85 ± 5.18 83.33 ± 6.40 355.8 ± 25.77 56.06 ± 3.95 53.68 ± 4.02 39.06 ± 4.00 18.87 ± 0.66 27.00 ± 3.71 19.94 ± 1.23 6.52 ± 0.95 4.13 ± 0.52 6.95 ± 0.90 5.86 ± 0.40 2.47 ± 0.62 634.0 138.6 45.87 818.5 1.72 ± 0.85 73.18 ± 1.55

254.0 ± 52.39 39.54 ± 10.29 41.49 ± 9.06 129.4 ± 20.03 760.6 ± 155.5 80.60 ± 3.64 109.8 ± 30.97 121.2 ± 45.43 24.08 ± 3.97 53.60 ± 15.34 18.93 ± 3.78 6.26 ± 0.90 4.42 ± 0.48 6.94 ± 0.95 5.96 ± 0.56 1.65 ± 0.71 1306 308.6 44.15 1658 3.56 ± 0.92 83.83 ± 2.22

550.4 ± 17.84 39.90 ± 9.01 36.10 ± 13.74 137.9 ± 51.04 634.3 ± 231.3 61.36 ± 18.38 74.99 ± 39.67 46.07 ± 34.77 24.39 ± 10.40 27.16 ± 10.67 23.48 ± 5.78 8.93 ± 4.21 5.33 ± 1.36 11.31 ± 3.61 9.36 ± 2.16 4.15 ± 1.63 1460 172.6 62.56 1695 5.07 ± 0.31 76.23 ± 3.58

608.6 ± 181.2 50.31 ± 4.26 46.52 ± 8.49 129.8 ± 14.82 655.7 ± 83.98 72.46 ± 14.02 71.05 ± 10.37 46.84 ± 10.95 21.99 ± 9.18 34.01 ± 11.69 25.03 ± 10.80 8.59 ± 2.06 7.97 ± 2.15 12.79 ± 1.24 11.95 ± 2.04 6.45 ± 9.23 1563 173.9 72.78 1810 8.63 ± 1.01 72.53 ± 1.29

586.3 ± 65.18 22.87 ± 9.03 40.50 ± 4.79 129.1 ± 25.93 885.6 ± 168.5 97.68 ± 43.47 145.7 ± 25.07 137.8 ± 33.36 19.35 ± 7.70 44.23 ± 24.36 17.93 ± 3.37 5.40 ± 1.58 3.98 ± 1.18 7.67 ± 2.50 6.76 ± 2.36 1.76 ± 0.59 1762 347.1 43.50 2153 3.33 ± 0.19 74.45 ± 5.26

The highest and the lowest ΣPAHs, L-PAHs and M-PAHs levels were found in the mantles and adductor, respectively (p < 0.01). But for HPAHs, the highest levels for were found in the hepatopancreas tissue, and the lowest were in the adductor and gills tissues (p < 0.01). No significant differences were found between gonads and gills of both bivalves (p > 0.05). This presumably depends on the different uptake and metabolic rates, which will be manifested among the five different tissues in bivalves. Accumulation of PAHs in mantles, hepatopancreas, gonads and gills could be also related to their lipid contents and physiological roles in metabolism, as described earlier. Mantles can accumulate the highest levels of PAHs possibly due to their direct contact with the ambient environments (Wang et al., 2005). Same as most organic pollutants, PAHs prefer to accumulate in certain tissues, such as in the hepatopancreas of invertebrates or in the liver of vertebrates (Dû-Lacoste et al., 2013; Lemaire et al., 1993). In the current study, hepatopancreas displayed 1.2 to 1.9 times higher HPAHs levels than those observed in all other tissues. The hepatopancreas is the major site for the digestion and processing of food, which can be expected to have the highest contents of H-PAHs in bivalves. Besides, the hepatopancreas is also a crucial xenobiotic metabolizing organ with high lipid contents, and higher levels of organic contaminants’ metabolites are usually found. Although metabolism in different organs in animals also plays an important role for accumulation of organic pollutants in their different body parts; however, for PAHs with low log Kow (< 5), such as Nap, Phe and Ant, no metabolism was observed in aquatic organisms (Thomann et al., 1992). A previous study has also reported that, in the liver of Parophrys vetulus, Nap (L-PAHs) was not as efficiently metabolized as BaP (H-PAHs) (Varanasi, Gmur, 1981). Thus, the high H-PAHs levels found in hepatopancreas in this study may be merely due to its high lipid contents, not related to discrepancies between metabolic rates of different PAHs in different tissues of bivalves. In comparison, the adductor, with no secretory or absorptive function and the worst lipid contents, contained lower levels of PAHs than those in other tissues. It may be a poor indicator for PAHs contamination in bivalves. Higher levels of PAHs in mantles and gills found in the current investigation may be interpreted as mainly due to their direct contact with the surrounding water resulting in greater and faster accumulation. 3.3. Correlations of PAH bioaccumulation in bivalve tissues As shown in Fig. S1, a radial analysis of the percentages for individual PAHs in different tissues may also clarify their transfer inside bivalves. Among the sixteen PAHs tested for, the distribution profiles of PAHs in both bivalves was remarkably uniform among all tissues analyzed, which suggests that the transfer of the PAHs in oysters and in mussels shares a similar pattern. In both bivalves, the PAH distributions were predominated by Phe (36.22–50.37% of total PAHs), followed by Nap, Flua and Pyr (13.82–33.62%, 3.93–9.67% and 2.59–7.31%, respectively) in most of the tissues, but this was not the case for adductor and gills of the oyster. Similar distribution patterns of 2–3 rings PAHs in bivalves are in accordance with a lot of earlier studies (Barhoumi et al., 2016; Olenycz et al., 2015; Sun et al., 2018). The PAHs other than 2–3 rings comprised less than 5% of the total. The dominance of Phe is consistent with the PAHs composition found in bivalves from other marine systems (Barhoumi et al., 2016; Mix,Schaffer, 1983). On the other hand, BghiP was the least abundant (0.08–0.4%), lower than BaP (0.2–0.6%). As benthic filter-feeding organisms, bivalves can filter and adsorb small particles through their gills. Their mantles and gills have direct contact with environmental media containing contaminants when their shells are open (Wang et al., 2005). Therefore, PAHs in bivalve mantles and gills were significantly correlated (r2 = 0.901, p < 0.0001) (Fig. S2), as well as their distribution pattern as above mentioned (Fig. S2). The highest correlation was found between hepatopancreas and gonads (r2 = 0.975, p < 0.0001), possibly due to their relatively higher lipid

Nap Ace Acy Flu Phe Ant Flua Pyr BaA Chy BbF BkF BaP InP DahA BghiP L-PAHs M-PAHs H-PAHs ΣPAHs Lipid (%) Water Content (%)

26.57 ± 12.54 15.24 ± 2.92 16.32 ± 4.82 67.77 ± 16.27 264.0 ± 71.9 41.02 ± 8.42 43.50 ± 11.21 41.82 ± 5.06 16.92 ± 2.53 18.41 ± 9.83 20.32 ± 1.33 6.43 ± 0.28 3.72 ± 0.37 6.55 ± 0.95 6.09 ± 0.78 2.41 ± 0.75 430.9 120.6 45.53 597.1 1.74 ± 0.18 76.82 ± 0.71

23.67 ± 20.24 24.00 ± 12.15 22.53 ± 11.76 94.05 ± 52.34 438.5 ± 267.8 49.82 ± 18.33 103.3 ± 30.91 152.2 ± 31.47 29.47 ± 12.56 81.66 ± 14.58 22.53 ± 7.68 5.52 ± 0.68 4.22 ± 1.96 7.62 ± 1.33 6.13 ± 0.55 2.20 ± 1.03 652.6 366.6 48.23 1067 3.92 ± 0.53 84.02 ± 0.64

324.5 ± 150.9 22.32 ± 15.42 22.41 ± 21.44 95.94 ± 14.92 464.6 ± 449.4 47.47 ± 23.25 72.24 ± 15.78 59.04 ± 19.83 30.16 ± 16.39 20.64 ± 17.02 22.79 ± 10.06 11.69 ± 10.02 5.02 ± 2.00 8.57 ± 4.93 6.16 ± 2.02 2.28 ± 0.22 977.2 182.1 56.52 1216 5.87 ± 1.9 76.33 ± 0.96

585.3 ± 98.54 31.47 ± 2.90 42.15 ± 7.93 143.9 ± 27.49 806.0 ± 181.6 65.41 ± 14.06 127.6 ± 59.04 120.3 ± 23.87 35.25 ± 14.49 40.19 ± 18.94 37.50 ± 16.57 12.19 ± 5.46 6.71 ± 2.43 12.72 ± 5.02 7.60 ± 2.01 3.21 ± 2.03 1674 323.4 79.92 2078 8.55 ± 0.89 75.68 ± 1.63

Adductor Mantles Hepatopancreas Adductor

Gills

Gonads

Perna viridis Pinctada martensii PAHs

Table 1 Levels of PAHs (ng g−1 dry weight) and lipid contents (%) in the different tissues of bivalves from the Li'an and Xincun Bays.

Gills

Gonads

Hepatopancreas

Mantles

H. Wang, et al.

5

Food and Chemical Toxicology 136 (2020) 111108

H. Wang, et al.

(2.78–4.25) found in the current study may be due to its low Kow value, relatively higher gill transfer efficiency, and dietary uptake rate. Lower BSAFs indicate higher biotransformation of PAHs in the two bivalve species. BSAF values less than 1 may suggest metabolism has occurred (Thorsen et al., 2004). In addition, lower BSAF values for pyrogenic PAHs should be considered less bioavailable (Thorsen et al., 2004). However, some other factors should be taken into consideration for evaluation and utilization of BSAFs, including differences in hydrophobic characteristics between bivalve's lipid contents and sediment organic carbons, lack of steady state between different environmental media, and other physiological conditions. Thus, cautions should be taken for interpretation of BSAFs. 3.5. PAHs source identification PAHs frequently occur in different environment compartments as complex mixtures and often result from the combination of petrogenic and pyrolytic sources. Pyrogenic PAHs are formed during high temperature and inadequate combustion processes, while petrogenic PAHs originate from spilled or leaked crude oil and/or their refined products. For source analysis, two different molecular diagnostic ratios (MDR), namely Ant/(Phe + Ant) and Flua/(Flua + Pyr), were widely used to identify the possible sources of PAHs in different environmental media (Soliman et al., 2014; Yunker et al., 2002). It is acknowledged that Ant/ (Phe + Ant) values more than 0.1 are related to pyrogenic sources and less than 0.1 are related to petroleum sources. For Flua/(Flua + Pyr), values more than 0.5 are related to combusted coal, wood, and/or grass, while values between 0.4 and 0.5 are related to combusted liquid fossil fuels and values less than 0.4 are related to petrogenic sources (Li et al., 2015b). In this study, we found that Ant/(Phe + Ant) ratios were higher than 0.1 in all of the sediment samples from Li'an and Xincun Bays, indicating significant pyrogenic sources (Fig. 3). It is also noted that Flua/(Flua + Pyr) values were higher than 0.5 in all the sediment samples, suggesting combusted coal/wood/grass origins. For seawater samples, however, the plot of Flua/(Flua + Pyr) against Ant/ (Phe + Ant) ratios indicates that PAHs were from mixed sources of pyrogenic and petrogenic. As for bivalves, as shown in Fig. 3, PAHs were from mixed sources of petrogenic and pyrogenic with primarily originated from combusted coal, wood, grass and fuels. The similarity between the MDRs of bivalves and ambient sediment indicates that the sources of the PAHs were probably the same, which was supported by their patterns of distribution and occurrence in both media. In summary, the results of MDR analysis revealed that the presence of combustion and unburned petroleum in the area studied, which could be linked to resident's or tourist's activities, terrigenous input and spills or leaks from boat machineries using petroleum products. Our data also suggest that most bivalves in this area may be exposed to and adsorbed a mixture of petrogenic and pyrogenic PAHs from both water columns and surrounding sediments.

Fig. 3. The mean isomeric ratios of PAH components in the bivalves at Li'an and Xincun bay.

contents (Table 1) and similar functions. The lipid content is one of the key factors influencing PAHs bioaccumulation in marine organisms (Sun et al., 2016). The lowest correlation was found between gills and hepatopancreas (r2 = 0.793, p < 0.0001), less than that between gills and gonads (r2 = 0.811, p < 0.0001). High correlation may also mean that the organs may have similar PAHs transition, metabolism and/or elimination. Nevertheless, based on our results, hepatopancreas appeared to be a promising tissue for evaluating H-PAHs levels in the two bivalve species studied. Not only did it contain relatively higher ΣPAHs and much higher H-PAHs concentrations than the others, but also its PAHs levels were in good correlation with those in the gonads (r2 = 0.975, p < 0.0001). On the other hand, the mantles can be considered as a good indicator tissue to reflect ΣPAHs in ambient environment because of the highest concentrations been detected there. Nevertheless, more field data and investigations on mechanisms of bioaccumulation in these animals should be warranted. 3.4. Biota-sediment accumulation factors PAHs in sediment play the important role in exposure and effects in aquatic organism. As bottom feeders, bivalves frequently can accumulate PAHs in sediment. In the current study, we have found that the levels of ΣPAH in the bivalves were significantly higher than those in the surrounding sediments (Table S7), which was in agreement with previous reports (Sun et al., 2016). It is also providing the evidence of PAHs bioaccumulation by bivalves through their surroundings. In order to elucidate the bioaccumulation patterns of individual PAHs in bivalve tissues, we correlated the log-transformed BSAFs of individual PAHs with their log Kow values (Luellen, Shea, 2002) for each type of bivalve tissue (Fig. S3). Based on our results, we found that the levels of PAHs in bivalve tissues were strongly negatively correlated with their Kow values (r2 = 0.501, p < 0.0001), indicating that the BSAFs decline with the increase of Kow values. The relationship between log BSAFs values in bivalve and log Kow values was also reflected by the role of the dietary route in bioaccumulation, which was consistent with previous studies (Liang et al., 2007; Sun et al., 2018). In their field study, Thomann and his group has estimated generic BSAFs as followed: 1.0–10.0 for organic compounds with log Kow values between 5.0 and 8.0, and 0.8–1.0 for those with log Kow values between 2.0 and 5.0 (Thomann et al., 1992). The present results revealed that the BSAFs of all PAHs were below the predicted values, except for that of Nap (Fig. S3). As a not readily metabolized organic compound, Nap with high BSAF values

3.6. Levels of BaP, ΣPAHs4 and ΣPAHs8 in bivalves The European Food Safety Authority (EFSA) has concluded that BaP is not a suitable marker for the occurrence of PAHs in food and that a system of four specific PAHs (ΣPAHs4 = BaP + BaA + BbF + Chy) or eight specific PAHs (ΣPAHs8 = ΣPAHs4 + BkF + BghiP + DahA + InP) would be the most suitable indicators of PAH in food, whilst maintaining a separate maximum level for BaP (European Commission, 2011). However, in our case, the concentrations of ΣPAHs4 and ΣPAHs8 in Pinctada martensii and Perna viridis samples (Fig. 4) showed no significant difference (p > 0.05). According to a scientific opinion of the former Scientific Committee on Food, BaP is used as a marker for the occurrence and effects of carcinogenic PAHs in food (European Commission, 2011), but BaP detected in the tissues of Pinctada martensii and Perna viridis with average concentration of 6

Food and Chemical Toxicology 136 (2020) 111108

H. Wang, et al.

Fig. 4. The concentrations of BaP, ΣPAHs4 and ΣPAHs8 in Pinctada martensii and Perna viridis.

1.08 μg kg−1 w w and 1.17 μg kg−1 w w (Fig. 4) also showed no significant difference (p > 0.05). The overall concentrations of ΣPAHs4 and ΣPAHs8 ranged between 13.76 and 28.33 μg kg−1 w w and 18.74–36.78 μg kg−1 w w, respectively (Table S3). The highest ΣPAHs4 and ΣPAHs8 were found in hepatopancreas of Pinctada martensii. BaP detected in the tissues of Pinctada martensii and Perna viridis ranged from 0.67 to 1.89 μg kg−1 w w (Table S3), and were similar to those found in other mollusc species (Conte et al., 2016; Hong et al., 2016; Uno et al., 2017). However, the observed concentrations of BaP and ΣPAHs4 in all samples were lower than the maximum residue level of bivalve mollusks (5 and 30.0 μg kg w. w) defined by the European Commission (European Commission, 2011).

body weight day−1. For both bivalve species, the estimated EDIs of ΣPAHs for consumption of adductor and gills were the lowest compared to those of other tissues. The highest values of EDIs were found for ingestion of hepatopancreas and mantles. THQ and CR from the ingestion of bivalves for the consumers with two different ingestion rates (IR of world and China averages) were shown in Table S8. By using the two IR values, ΣTHQs of tissue in the two bivalves calculated were 2.60 × 10−6-9.82 × 10−4 and 9.19 × 10−6-3.48 × 10−3, respectively (Fig. 5B). Nap, Flu and Pyr were the PAHs that contributed the most for ΣTHQs, and their THQs varied from 1.85 × 10−5-2.62 × 10−3, 3.68 × 10−5-3.28 × 10−4, and 3.16 × 10−5-4.58 × 10−4, respectively (Table S8). Nevertheless, ΣTHQ values were always below 1, a value considered as a presence of no potential risks (USEPA, 2017). These results also indicated the PAHs levels found in oysters or mussels may pose no immediately or longterm non-carcinogenic effects for most of the consumers. For carcinogenic risk, the calculated ΣCR values for the two scenarios, using either world or China average ingestion rate, were all higher than 10−6 (world: 1.71 × 10−6-4.20 × 10−6; China: 6.06 × 10−6-1.49 × 10−5) (Fig. 5C). Overall, the highest contributors for ΣTR is DahA and BaP, which is similiar with the findings by other studies (Oliveira et al., 2018; Falcó et al., 2005). The ΣTR caused by exposure to PAHs is higher than the acceptable risk level (10−6), but lower than the priority risk level (10−4) proposed by USEPA (USEPA, 1989). Therefore, it is presumed that the ingestion of the selected bivalves may pose concerns of carcinogenic risks for consumers with higher consumption rate. Our results was different from Yu's report, in which his group calculated cancer rates for ingestion of oysters from Basuo Habor, Hainan, were exceeded 10−5 (Yu et al., 2016). Nevertheless, with regards to risk assessment, the risks are totally associated with both the levels of contaminates in the food (bivalves in this study) and the daily intake rate of the food. In the case of this study, the risk in certain population, if any, may be attributed to higher consumption rate of bivalves.

3.7. Potential health risks for consumers Estimate of EDI for PAHs is important due to several epidemiological studies showed that colorectal, gastric, and gastrointestinal tract cancers have been associated with dietary exposure to PAHs (Diggs et al., 2011; Liao et al., 2014; Wang, 2008). The calculated EDIs for individual PAH using average consumption rates of world (2.50 kg per capita per year) and China (8.85 kg per capita per year) are showed in Table S8, while for ΣPAHs, they are presented in Fig. 5A. In this study, higher intakes were found for the 3-ring and 4-ring PAHs. It should be noted that the average consumption rate for Chinese consumers is almost three-fold higher than those reported for consumers in Samoa, Australia, Brunei Darussalam and New Zealand (3.08, 2.92, 2.78, 1.93 kg per capita per year), and is comparable to the levels for Spain or Republic of Korea consumers (7.1 and 9.61 kg per capita per year) (FAO, 2019). In this study, the estimated EDIs for BaP and ΣPAHs varied from 7.81 × 10−5-4.38 × 10−4 and 3.78 × 10−21.59 × 10−1 μg kg−1 body weight day−1 for ingestion or the whole organism, and ranged between 6.60 × 10−5-6.56 × 10−4 and 1.35 × 10−2-1.91 × 10−1 μg kg−1 body weight day−1 for ingestion of different tissues of oysters and mussels. Regarding consumers in China, the EDI of BaP and ΣPAHs were almost 4 times higher than populations using the world average consumption rate, and mean were 3.57 × 10−4 and 1.47 × 10−1 μg kg−1 body weight day−1 for consumption of the whole organism. These levels were also higher than those reported for ingestion of green ormer Haliotis tuberculata (Linnaeus, 1758) in Italian population (Conte et al., 2016), or for sand dwelling mussel (Donax trunculus) in Mediterranean general population (Ferrante et al., 2018). However, EDI of BaP is below the acceptable daily intake of 10 ng−1 kg−1 day suggested by FAO/WHO (JECFA, 2019). In the case of ingestion of different tissues of oysters and mussels, the calculated EDIs varied from 2.42 × 10−5-1.91 × 10−1 μg kg−1

4. Conclusions The occurrence of PAHs in the seawater, sediment and two bivalve species from the Li'an and Xincun Bays in Hainan Island was investigated. Significant differences between M-PAHs and H-PAHs levels in the two bivalves were observed. However, for the two bivalves examined, very similar PAHs patterns in tissues were observed in which LPAHs mainly predominated, followed by M-PAHs. These were similar to those found in the seawater. Analysis on the ratios of specific PAH compounds showed that the pyrolytic sources played a significant role in the pollution of the local coastal environment. Because of higher 7

Food and Chemical Toxicology 136 (2020) 111108

H. Wang, et al.

Fig. 5. Estimated daily intake of PAHs (A), the target hazard quotient (B) and carcinogenic risks (C) by the ingestion of Pinctada martensii and Perna viridis.

Transparency document

ΣPAHs concentrations detected in the mantles and hepatopancreas, they can be considered as indicator tissues for environmental monitoring in the future. There was a significantly negative correlation between individual PAHs' BSAFs and their corresponding log Kow, which may be interpreted by higher transfer efficiencies for PAHs with lower Kow via gills. The EDI, THQ and CR were estimated. Although the results showed no potential non-carcinogenic effects is present, carcinogenic risks may exist for certain population of heavier consumption of bivalves. Nevertheless, based on the precautionary principle, in order to protect the health of general publics, it is crucial to confirm the PAHs pollution sources as well as to minimize the contamination and their associated risks.

Transparency document related to this article can be found online at https://doi.org/10.1016/j.fct.2019.111108. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.fct.2019.111108. References Abdel-Shafy, H.I., Mansour, M.S.M., 2016. A review on polycyclic aromatic hydrocarbons: source, environmental impact, effect on human health and remediation. Egypt. J. Petrol. 25, 107–123. Ashraful, M.A.K., Assim, Z.B., Ismail, N., 2009. Food and feeding biology of green mussel, Perna viridis from the Bay of Bengal coast, Bangladesh. Ecol. Environ. Conserv. 15, 415–420. Barhoumi, B., Megdiche, Y.E., Clérandeau, C., Ameur, W.B., Mekni, S., Bouabdallah, S., Derouiche, A., Touil, S., Cachot, J., Driss, M.R., 2016. Occurrence of polycyclic aromatic hydrocarbons (PAHs) in mussel (Mytilus galloprovincialis) and eel (Anguilla anguilla) from Bizerte lagoon, Tunisia, and associated human health risk assessment. Cont. Shelf Res. 124, 104–116. Chen, Y., Zhu, L., Zhou, R., 2007. Characterization and distribution of polycyclic aromatic hydrocarbon in surface water and sediment from Qiantang River, China. J. Hazard Mater. 141, 148–155. Chen, H., Song, Q., Diao, X., Zhou, H., 2016. Proteomic and metabolomic analysis on the toxicological effects of Benzo[a]pyrene in pearl oyster Pinctada martensii. Aquat. Toxicol. 175, 81–89. Conte, F., Copat, C., Longo, S., Conti, G.O., Grasso, A., Arena, G., Dimartino, A., Brundo, M.V., Ferrante, M., 2016. Polycyclic aromatic hydrocarbons in Haliotis tuberculata (Linnaeus, 1758) (Mollusca, Gastropoda): considerations on food safety and source investigation. Food Chem. Toxicol. 94, 57–63. Cui, L., Ge, J., Zhu, Y., Yang, Y., Wang, J., 2015. Concentrations, bioaccumulation, and human health risk assessment of organochlorine pesticides and heavy metals in edible fish from Wuhan, China. Environ. Sci. Pollut. Res. Int. 22, 15866–15879. Deeb, K.Z.E., Said, T.O., Naggar, M.H.E., Shreadah, M.A., 2007. Distribution and sources of aliphatic and polycyclic aromatic hydrocarbons in surface sediments, fish and bivalves of abu qir bay (Egyptian Mediterranean sea). Bull. Environ. Contam. Toxicol. 78, 373–379. Diggs, D.L., Huderson, A.C., Harris, K.L., Myers, J.N., Banks, L.D., Rekhadevi, P.V., Niaz, M.S., Ramesh, A., 2011. Polycyclic aromatic hydrocarbons and digestive tract

CRediT authorship contribution statement Haihua Wang: Conceptualization, Investigation, Formal analysis, Writing - original draft. Wei Huang: Conceptualization, Investigation. Ying Gong: Formal analysis. Chienmin Chen: Writing - review & editing. Tengyun Zhang: Investigation. Xiaoping Diao: Conceptualization, Supervision, Funding acquisition. Declaration of competing interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgements The present work was supported in part by the National Natural Science Foundation of China (No. 31560165) and the Department of Science and Technology of Hainan Province (ZDYF2018122). 8

Food and Chemical Toxicology 136 (2020) 111108

H. Wang, et al. cancers: a perspective. J. Environ. Sci. Health Part C Environ. Carcinog. Ecotoxicol. Rev. 29, 324–357. Le Dû-Lacoste, M., Akcha, F., Dévier, M.H., Morin, B., Burgeot, T., Budzinski, H., 2013. Comparative study of different exposure routes on the biotransformation and genotoxicity of PAHs in the flatfish species, Scophthalmus maximus. Environ. Sci. Pollut. Res. 20, 690–707. European Commission, 2011. Commission Regulation (EU) No 835/2011 of 19 August 2011 Amending Regulation (EC) No 1881/2006 as Regards Maximum Levels for Polycyclic Aromatic Hydrocarbons in Foodstuffs. Falcó, G., Bocio, A., Llobet, J.M., Domingo, J.L., 2005. Health risks of dietary intake of environmental pollutants by elite sportsmen and sportswomen. Food Chem. Toxicol. An Int. J. Publ. Br. Ind. Biol. Res. Assoc. 43, 1713–1721. FAO, 2016. Food and agriculture organization of the United Nations. Available at: http:// faostat.fao.org/site/354/default.aspx. FAO, 2019. Fishery statistical collections, fishery and aquaculture department. National aquaculture sector overview. Food Agric. Organ. U. N Available at:, Accessed date: May 2019. http://www.fao.org/fishery/countrysector/naso_china/en. Ferrante, M., Zanghì, G., Cristaldi, A., Copat, C., Grasso, A., Fiore, M., Signorelli, S.S., Zuccarello, P., Oliveri Conti, G., 2018. PAHs in seafood from the Mediterranean Sea: an exposure risk assessment. Food Chem. Toxicol. 115, 385–390. Hong, W., Jia, H., Li, Y., Sun, Y., Liu, X., Wang, L., 2016. Polycyclic aromatic hydrocarbons (PAHs) and alkylated PAHs in the coastal seawater, surface sediment and oyster from Dalian, Northeast China. Ecotoxicol. Environ. Saf. 128, 11–20. IRISintegrated risk information system Database developed and maintained by U.S. Environmental protection agency. Available at: https://cfpub.epa.gov/ncea/iris_ drafts/atoz.cfm?list_type=alpha Accessed date: 15 May 2019. JECFA, 2019. Evaluations of the joint FAO/WHO expert committee on food additives. Available at: http://apps.who.int/food-additives-contaminants-jecfa-database/ chemical.aspx?chemID=4306. Kaushik, C., Haritash, A.K., 2006. Polycyclic Aromatic Hydrocarbons (PAHs) and Environmental Health. Kim, K.H., Jahan, S.A., Kabir, E., Brown, R.J., 2013. A review of airborne polycyclic aromatic hydrocarbons (PAHs) and their human health effects. Environ. Int. 60, 71–80. Ko, F.C., Chang, C.W., Cheng, J.O., 2014. Comparative study of polycyclic aromatic hydrocarbons in coral tissues and the ambient sediments from Kenting National Park, Taiwan. Environ. Pollut. 185, 35. Kofi, E.D., Roberta, A., Joseph, A., Akwansah, E.G., Dodoo, D.K., 2005. Seasonal variation of polycyclic aromatic hydrocarbon (PAH) contamination in (oysters) and sediments in three Ghanaian coastal ecosystems. Res. J. Environ. Sci. 12, 63–72. Law, R.J., Dawes, V.J., Woodhead, R.J., Matthiessen, P., 1997. Polycyclic aromatic hydrocarbons (PAH) in seawater around England and Wales. Mar. Pollut. Bull. 34, 306–322. Lemaire, P., Den Besten, P.J., O'Hara, S.C.M., Livingstone, D.R., 1993. Comparative metabolism of benzo[a]pyrene by microsomes of hepatopancreas of the shore crab Carcinus maenas L. and digestive gland of the common mussel Mytilus edulis L. Ploy. Arom. Comp. 3, 1133–1140. Li, P., Cao, J., Diao, X., Wang, B., Zhou, H., Han, Q., Zheng, P., Li, Y., 2015. Spatial distribution, sources and ecological risk assessment of polycyclic aromatic hydrocarbons in surface seawater from Yangpu Bay, China. Mar. Pollut. Bull. 93, 53–60. Li, Y., Li, P., Ma, W., Song, Q., Zhou, H., Han, Q., Diao, X., 2015. Spatial and temporal distribution and risk assessment of polycyclic aromatic hydrocarbons in surface seawater from the Haikou Bay, China. Mar. Pollut. Bull. 92, 244–251. Liang, Y., Tse, M.F., Young, L., Wong, M.H., 2007. Distribution patterns of polycyclic aromatic hydrocarbons (PAHs) in the sediments and fish at Mai Po Marshes Nature Reserve, Hong Kong. Water Res. 41, 1303–1311. Liao, L.M., Hofmann, J.N., Farin, K., Strickland, P.T., Bu-Tian, J., Gong, Y., Hong-Lan, L., Nathaniel, R., Wei, Z., Wong-Ho, C., 2014. Polycyclic aromatic hydrocarbons and risk of gastric cancer in the Shanghai Women's Health Study. Int. J. Mol. Epidemiol. Genet. 5, 140–144. Logan, D.T., 2007. Perspective on ecotoxicology of PAHs to fish. Hum. Ecol. Risk Assess. Int. J. 13, 302–316. Loret, P., Annie, P., Cedric, B., Delesalle, B., 2000. Phytoplankton composition and selective feeding of the pearl oyster Pinctada margaritifera in the Takapoto lagoon (Tuamotu Archipelago, French Polynesia): in situ study using optical microscopy and HPLC pigment analysis. Mar. Ecol. Prog. 199, 55–67. Luellen, D.R., Shea, D., 2002. Calibration and field verification of semipermeable membrane devices for measuring polycyclic aromatic hydrocarbons in water. Environ. Sci. Technol. 36, 1791–1797. Mashroofeh, A., Bakhtiari, A.R., Pourkazemi, M., 2015. Distribution and composition pattern of polycyclic aromatic hydrocarbons in different tissues of sturgeons collected from Iranian coastline of the Caspian Sea. Chemosphere 120, 575–583. Mix, M.C., Schaffer, R.L., 1983. Concentrations of unsubstituted polynuclear aromatic hydrocarbons in bay mussels (Mytilus edulis) from Oregon, USA. Mar. Environ. Res. 9, 193–209. Montuori, P., Aurino, S., Garzonio, F., Sarnacchiaro, P., Nardone, A., Triassi, M., 2016. Distribution, sources and ecological risk assessment of polycyclic aromatic hydrocarbons in water and sediments from Tiber River and estuary. Italy. Sci. Total Environ. 566, 1254–1267. Moslen, M., Miebaka, C.A., Boisa, N., 2019. Bioaccumulation of Polycyclic Aromatic Hydrocarbon (PAH) in a Bivalve (Arca senilis- blood cockles) and health risk assessment. Toxicol. Rep. 6. OECD-FAO, 2019. Organization for Economic Cooperation and Development – Food and Agriculture Organization. Olenycz, M., Sokołowski, A., Niewińska, A., Wołowicz, M., Namieśnik, J., Hummel, H., Jansen, J., 2015. Comparison of PCBs and PAHs levels in European coastal waters

using mussels from the Mytilus edulis complex as biomonitors. Oceanologia 57, 196–211. Oliveira, M., Gomes, F., Torrinha, Á., Ramalhosa, M.J., Delerue-Matos, C., Morais, S., 2018. Commercial octopus species from different geographical origins: levels of polycyclic aromatic hydrocarbons and potential health risks for consumers. Food Chem. Toxicol. 121, 272–282. Pan, Y., Wang, Y.X., Li, S., Chen, Y.J., Chong, L., Huang, L.L., Lu, W.Q., Qiang, Z., 2017. Urinary metabolites of polycyclic aromatic hydrocarbons, sperm DNA damage and spermatozoa apoptosis. J. Hazard Mater. 329, 241–248. Ranjbar, A.J., Riyahi, A.B., Aliabadian, M., Shadmehri, A.T., 2017. Spatial distribution and composition of aliphatic hydrocarbons, polycyclic aromatic hydrocarbons and hopanes in superficial sediments of the coral reefs of the Persian Gulf, Iran. Environ. Pollut. 224, 195–223. Rengarajan, T., Rajendran, P., Nandakumar, N., Lokeshkumar, B., Rajendran, P., Nishigaki, I., 2015. Exposure to polycyclic aromatic hydrocarbons with special focus on cancer. Asian Pac. J. Trop. Biomed. 5, 182–189. Richardson, B.J., Mak, E., De Luca-Abbott, S.B., Martin, M., Mcclellan, K., Lam, P.K., 2008. Antioxidant responses to polycyclic aromatic hydrocarbons and organochlorine pesticides in green-lipped mussels (Perna viridis): do mussels "integrate" biomarker responses? Mar. Pollut. Bull. 57, 503–514. Soliman, Y.S., Ansari, E.M.S.A., Wade, T.L., 2014. Concentration, composition and sources of PAHs in the coastal sediments of the exclusive economic zone (EEZ) of Qatar, Arabian Gulf. Mar. Pollut. Bull. 85, 542–548. Song, Q., Zhou, H., Han, Q., Diao, X., 2017. Toxic responses of Perna viridis hepatopancreas exposed to DDT, benzo(a)pyrene and their mixture uncovered by iTRAQbased proteomics and NMR-based metabolomics. Aquat. Toxicol. 192, 48–57. Soriano, J.A., Viñas, L., Franco, M.A., González, J.J., Ortiz, L., Bayona, J.M., Albaigés, J., 2006. Spatial and temporal trends of petroleum hydrocarbons in wild mussels from the Galician coast (NW Spain) affected by the Prestige oil spill. Sci. Total Environ. 370, 80–90. Sun, R.X., Lin, Q., Ke, C.L., Du, F.Y., Gu, Y.G., Cao, K., Luo, X.J., Mai, B.X., 2016. Polycyclic aromatic hydrocarbons in surface sediments and marine organisms from the Daya Bay, South China. Mar. Pollut. Bull. 103, 325–332. Sun, R., Sun, Y., Li, Q.X., Zheng, X., Luo, X., Mai, B., 2018. Polycyclic aromatic hydrocarbons in sediments and marine organisms: implications of anthropogenic effects on the coastal environment. Sci. Total Environ. 640, 264–272. Tam, N.F.Y., Ke, L., Wang, X.H., Wong, Y.S., 2001. Contamination of polycyclic aromatic hydrocarbons in surface sediments of mangrove swamps. Environ. Pollut. 114, 255–263. Thomann, R.V., Connolly, J.P., Parkerton, T.F., 1992. An equilibrium model of organic chemical accumulation in aquatic food webs with sediment interaction. Environ. Toxicol. Chem. 11, 615–629. Thorsen, W.A., Cope, W.G., Shea, D., 2004. Bioavailability of PAHs: effects of soot carbon and PAH source. Environ. Sci. Technol. 38, 2029–2037. Tomaru, Y., Kawabata, Z., Nakano, S., 2000. Consumption of picoplankton by the bivalve larvae of Japanese pearl oyster Pinctada fucata martensii. Mar. Ecol. Prog. 192, 195–202. Tu, Z.G., Han, T.S., Chen, X.H., Rui, W.U., Wang, D.R., 2016. Thecommunity structure and diversity of macrobenthos in Linshui Xincungang and Li'angang seagrass special protected Area,Hainan. Mar. Environ. Sci. 35, 41–48. Uno, S., Kokushi, E., Añasco, N.C., Iwai, T., Ito, K., Koyama, J., 2017. Oil spill off the coast of Guimaras Island, Philippines: distributions and changes of polycyclic aromatic hydrocarbons in shellfish. Mar. Pollut. Bull. 124, 962–973. USEPA, 1989. Risk assessment guidance for superfund. Human Health Evaluation Manual, vol. I Office of Emergency and Remedial Response, Washington DC EPA/ 540/1–89/002. USEPA, 1993. Provisional Guidance for Quantitative Risk Assessment of Polycyclic Aromatic Hydrocarbons. EPA/600/R-93/089. USEPA, 2017. Risk based concentration table. Available at: https://www.epa.gov/sites/ production/files/2016-06/documents/master_sl_table_run_may 2016.pdf Accessed date: May 2019. Valavanidis, A., Vlachogianni, T., Triantafillaki, S., Dassenakis, M., Androutsos, F., Scoullos, M., 2008. Polycyclic aromatic hydrocarbons in surface seawater and in indigenous mussels (Mytilus galloprovincialis) from coastal areas of the Saronikos Gulf (Greece). Estuar. Coast Shelf Sci. 79, 733–739. Varanasi, U., Gmur, D.J., 1981. Hydrocarbons and metabolites in English sole (Parophrys vetulus) exposed simultaneously to [3H]benzo[a]pyrene and [14C]naphthalene in oilcontaminated sediment. Aquat. Toxicol. 1, 49–67. Vieweg, I., Hop, H., Brey, T., Huber, S., Jr, A.W., Gabrielsen, G.W., 2012. Persistent organic pollutants in four bivalve species from Svalbard waters. Environ. Pollut. 161, 134–142. Vives, I., Grimalt, J.O., Fernández, P., Rosseland, B., 2004. Polycyclic aromatic hydrocarbons in fish from remote and high mountain lakes in Europe and Greenland. Sci. Total Environ. 324, 67–77. Wang, J., 2008. Carcinogen metabolism genes, meat intake, and colorectal cancer risk. 130, 1898–1907 Dissertations & Theses - Gradworks. Wang, S.H., Wang, X.H., Hong, H.S., 2005. Specific characteristic of tissue bioaccumulation of PAHs in green lipped mussels (Perna Viridis) at different growth stages. Mar. Environ. Sci. 24, 29–32. Wang, Y., Wang, T., Li, A., Fu, J., Wang, P., Zhang, Q., Jiang, G., 2008. Selection of bioindicators of polybrominated diphenyl ethers, polychlorinated biphenyls, and organochlorine pesticides in mollusks in the Chinese Bohai sea. Environ. Sci. Technol. 42, 7159–7165. Wang, L., Pan, L., Liu, N., Liu, D., Xu, C., Miao, J., 2011. Biomarkers and bioaccumulation of clam Ruditapes philippinarum in response to combined cadmium and benzo[α] pyrene exposure. Food Chem. Toxicol. An Int. J. Publ. Br. Ind. Biol. Res. Assoc. 49,

9

Food and Chemical Toxicology 136 (2020) 111108

H. Wang, et al. 3407–3417. Wang, D.Q., Yu, Y.X., Zhang, X.Y., Zhang, S.H., Pang, Y.P., Zhang, X.L., Yu, Z.Q., Wu, M.H., Fu, J.M., 2012. Polycyclic aromatic hydrocarbons and organochlorine pesticides in fish from Taihu Lake: their levels, sources, and biomagnification. Ecotoxicol. Environ. Saf. 82, 63–70. Wang, H., Cui, L., Cheng, H., Zhang, Y., Diao, X., Wang, J., 2017. Comparative studies on the toxicokinetics of benzo[a]pyrene in Pinctada martensii and Perna viridis. Bull. Environ. Contam. Toxicol. 98, 1–7. Wei, Q., Sun, P., Xu, Z., Wang, Z., Zang, J., 2010. Shellfish aquaculture environment safety evaluation and monitoring system technology research. Ocean Dev. Manag. 3, 48–51. Weinberg, J.R., Leavitt, D.F., Lancaster, B.A., Capuzzo, J.M., 1997. Experimental field studies with Mya arenaria (Bivalvia) on the induction and effect of hematopoietic

neoplasia. J. Invertebr. Pathol. 69, 183–194. White, P.A., 2002. The genotoxicity of priority polycyclic aromatic hydrocarbons in complex mixtures. Mutat. Res. Genet. Toxicol. Environ. Mutagen. 515, 85–98. Xie, J., Zhao, C., Han, Q., Zhou, H., Li, Q., Diao, X., 2017. Effects of pyrene exposure on immune response and oxidative stress in the pearl oyster, Pinctada martensii. Fish Shellfish Immunol. 63, 237–244. Yu, Z., Lin, Q., Gu, Y., Ke, C., Sun, R., 2016. Spatial–temporal trend and health implications of polycyclic aromatic hydrocarbons (PAHs) in resident oysters, South China Sea: a case study of Eastern Guangdong coast. Mar. Pollut. Bull. 110, 203–211. Yunker, M.B., Macdonald, R.W., Vingarzan, R., Mitchell, R.H., Goyette, D., Sylvestre, S., 2002. PAHs in the Fraser River basin: a critical appraisal of PAH ratios as indicators of PAH source and composition. Org. Geochem. 33, 489–515.

10