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Distribution of alkylphenols in the Pearl River Delta and adjacent northern South China Sea, China
Chemosphere 63 (2006) 652–661 www.elsevier.com/locate/chemosphere
Distribution of alkylphenols in the Pearl River Delta and adjacent northern South China Sea, China Bing Chen a,b, Jing-Chun Duan a, Bi-xian Mai a,*, Xiao-Jun Luo a, Qing-Shu Yang a, Guo-Ying Sheng a, Jia-Mo Fu a a
State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, P.O. Box 1130, Guangzhou 510640, PR China b School of Environmental Science and Engineering, South China University of Technology, Guangzhou 510641, PR China Received 17 March 2005; received in revised form 26 July 2005; accepted 8 August 2005 Available online 10 October 2005
Abstract The occurrence of alkylphenols (APs) was investigated in surface water and sediments from the Pearl River Delta and adjacent northern South China Sea. Most of the water samples contained detectable amounts of APs, ranging up to 0.628 lg l1 for nonylphenol (NP) and 0.068 lg l1 for octylphenol (OP). APs were found in all of the sediment samples with concentrations ranging from 59 to 7808 lg kg1 for NP and from 1 to 93 lg kg1 for OP. The Zhujiang River showed the highest concentrations of APs in both water and sediments. Significant decrease of APs concentrations going from the Zhujiang River to the Shiziyang River was observed. The Xijiang River contained concentrations of APs slightly higher in water but relatively lower in sediments than the Lingding Bay, which might be attributed to their different hydrodynamic and sedimentary characteristics. There was a decreasing trend of APs in water from the rivers to the estuary and further to the sea on the whole. In the Lingding Bay and its outer waters, concentrations of APs in sediments increased to a maximum and then decrease seaward, which was consistent with the distribution trend of the sediment organic carbon contents. Linear regression analyses showed the concentrations of APs were markedly correlated with the sediment organic carbon contents, indicating that the sediment organic carbon is an important factor controlling the levels of APs in sediments. 2005 Elsevier Ltd. All rights reserved. Keywords: Alkylphenol; Water; Sediment; Pearl River Delta; South China Sea
1. Introduction Alkylphenol ethoxylates (APEs) are the worldÕs third largest group of surfactants with annual global produc* Corresponding author. Tel.: +86 20 85290146; fax: +86 20 85290706. E-mail addresses: [email protected] (B. Chen), nancymai@ gig.ac.cn (B.-x. Mai).
tion of 500 000 tons (Renner, 1997). They have been widely used in domestic detergents, pesticide formulations and industrial products, such as textiles, coatings, paints, lube oils, fuels, metals, plastics, pulp and paper. Nonylphenol ethoxylates (NPEs) and octylphenol ethoxylates (OPEs) are two of the most common surfactants in the marketplaces. NPEs account for about 80% of APEs, and OPEs account for most of the remainder (Renner, 1997). APEs enter the aquatic environment
0045-6535/$ - see front matter 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2005.08.004
B. Chen et al. / Chemosphere 63 (2006) 652–661
primarily via industrial and domestic wastewater with or without treatment due to their application characteristics. They can be biodegraded to lower ethoxymer and subsequently generate more persistent and lipophilic APs under anaerobic conditions (Thiele et al., 1997; Ying et al., 2002). There has been increasing concern over the presence of APs in the environment because of their toxicity (McLeese et al., 1981; Comber et al., 1993) and estrogenic activity (Soto et al., 1991; Jobling and Sumpter, 1993; Jobling et al., 1996). Since the route of wildlife and human exposure to these substances is mainly through water, the occurrence of APs has been investigated in a variety of rivers, lakes, estuaries and coastal waters around the world (e.g. Naylor et al., 1992; Bennie et al., 1997; Blackburn et al., 1999; Tsuda et al., 2000; Bester et al., 2001; Ferguson et al., 2001; Isobe et al., 2001; Jonkers et al., 2003; Rice et al., 2003; Basheer et al., 2004; Li et al., 2004). Reported levels in surface water ranged from <0.01 to 644 lg l1 for NP and from <0.005 to 0.18 lg l1 for OP (Bennie et al., 1997; Sole et al., 2000; Isobe et al., 2001). APs in water tend to be associated with particulate matter due to their hydrophobic nature. They are transported and finally deposited to the bottom sediments (Ferguson et al., 2001; Isobe et al., 2001). Relative to their short life in water, APs tend to accumulate over long periods of time in sediments (Liber et al., 1999; Shang et al., 1999). Levels of APs found in surface sediments were therefore substantially higher than the aqueous levels. Their concentrations ranged from <0.4 to 72 000 lg kg1 for NP and from <0.5 to 1800 lg kg1 for OP (Bennie et al., 1997; Bolz et al., 2001; Jonkers et al., 2003). Although distribution of APs in the aquatic environment has been documented in many countries, little information on APs is available in China where the production of APEs is over 50 000 tons annually (Huang, 1998). The Pearl River is the second largest river in China. It has a great number of tributaries and streams, forming a complicated watershed called the Pearl River Delta (Fig. 1). Approximately 3.3 · 1011 m3 of freshwater with suspended solids of 7.1 · 107 tons annually flow into the Pearl River Estuary via eight major outlets (Xu et al., 1993), and ultimately into the northern part of South China Sea. The Pearl River Delta covers an area of 41 700 km2 with a population of 41 million. As a result of rapid economic development in recent decades, the Pearl River Delta has become one of the most prosperous regions in China. However, the high growth of population, the rapid development of industry and agriculture, and the use of chemicals have caused serious pollution problems, leading to excessive release of pollutants into this region. It is estimated that industrial and domestic wastewater amounted to 3.0 · 109 m3 per year in the Pearl River Delta (Liu and Wu, 2004). Recent studies have revealed high levels of persistent organic
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pollutants and heavy metals in the river and estuarine environments (Cheung et al., 2003; Fu et al., 2003). To date, there has not been any report concerning APs contamination in this region. In this study, the occurrence of APs in water and sediments from the Pearl River Delta and adjacent northern South China Sea has been investigated. In addition, the distribution of APs in the aquatic environment combining rivers and estuary with sea as a whole has been discussed, which is important in assessing the impact of anthropogenic activities on the aquatic ecosystem and designing the management and conservation policies of this region.
2. Materials and methods 2.1. Chemicals 4-nonylphenol (mixture of compounds with branched side chain), 4-tert-octylphenol (93%), 4-tert-butylphenol (98%) and 4-a-cumylphenol (98%) were purchased from Tokyo Chemical Industry (Tokyo, Japan). Organic solvents were analytical grade and redistilled using a glass system. Neutral silica gel (80–100 mesh) was Soxhlet extracted with dichloromethane for 48 h, activated at 180 C for 12 h and then deactivated by adding 5% distilled water. 2.2. Study area and sample collection Three major tributaries of the Pearl River within the Pearl River Delta (Fig. 1) were selected for sampling. The first tributary comprises the Zhujiang River (ZR1– ZR5) and Shiziyang River (SR1–SR6). The Zhujiang River is situated in the northeast of the Pearl River Delta. It runs through the city of Guangzhou, the highly urbanized and densely populated provincial capital (7 million people) with numerous factories and industries, and are expected to be heavily contaminated with pollutants of industrial and domestic wastewater. The Shiziyang River receives inflows from the Zhujiang River and Dongjiang River, and empties into the Lingding Bay through the outlet of Humen. The second tributary is the Xijiang River (XR1–XR8), lying in the southwest of the Pearl River Delta. It runs across less developed areas and drains into the estuary through the Modaomen, Jitimen, and Hutiaomen outlets. The last tributary is the Dongjiang River (DR1–DR6), located in the northeast of the Pearl River Delta. It run through the city of Dongguan, where manufacturing industry is well developed, and flows into the Shiziyang River via Northernkou, Mayongkou, Daoyunhaikou, and Southernkou outlets. The Pearl River Estuary is one of the largest estuaries in the world (Fig. 1). The coast of Macao (MC1–MC4), located in the west side of the estuary, was selected for sampling. It is a known depositional zone for persistent
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Fig. 1. Map showing sampling sites in the Pearl River Delta and adjacent northern South China Sea. The numbers (1–8) indicate the eight major outlets to the South China Sea: (1) Humen; (2) Jiaomen; (3) Honqilimen; (4) Hengmen; (5) Modaomen; (6) Jitimen; (7) Hutiaomen; (8) Yamen. The numbers (9–12) indicate the four outlets to the Shiziyang River: (9) Northernkou; (10) Mayongkou; (11) Daoyunhaikou; (12) Southernkou. ZR, Zhujiang River; SR, Shiziyang River; XR, Xijiang River; DR, Dongjiang River; MC, Macao Coast; LB, Lingding Bay; SS, South China Sea.
organic pollutants in the Pearl River Delta. Relatively high levels of PCBs, DDTs and PAHs were found in surface and core sediments from Macao Harbor (Kang et al., 2000; Mai et al., 2002; Zhang et al., 2002; Mai et al., 2003). The main Pearl River Estuary, i.e. Lingding Bay (LB1–LB5), and its outer waters (SS1–SS4) were also selected for sampling. Lying in the northern part of the estuary, the Lingding Bay is funnel shaped and receives runoff from four outlets, i.e. Humen, Jiaomen,
Hongqilimen, and Hengmen. In addition, it is also greatly impacted by input from Macao and Hong Kong, one of the most populated regions in the world. Therefore, the Lingding Bay is considered an important pathway for pollutant transport into the South China Sea. The South China Sea is the largest marginal sea next to the western boundary of the Pacific Ocean. The water quality of the northern South China Sea, except the area near the estuary, achieved the second grade of National
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the cartridge with 20 ml methanol. The sediments were freeze-dried and homogenized. Each sample (10 g) was spiked with 4-tert-butylphenol as a surrogate standard and Soxhlet extracted with dichloromethane for 48 h. All the obtained extracts were concentrated and solvent-exchanged to hexane. The hexane extracts were purified through silica gel column (1 cm i.d. · 9 cm). The column was eluted with 20 ml dichloromethane/ hexane (4:6) and 40 ml dichloromethane/hexane (8:2). The later was collected as APs fraction and concentrated to 0.1 ml for water samples and 0.5 ml for sediment samplers under a gentle N2 stream. 4-a-Cumylphenol as internal standard was added to the sample prior to GC–MS analysis. APs were analyzed on HP5890 Series II gas chromatograph coupled with HP5972 mass selective detector in selected ion monitoring mode with DB-5 fused silica capillary column (30 m · 0.25 mm i.d., 0.25 lm film thickness). The injector and detector temperatures were set at 280 C and 300 C, respectively. The oven temperature program was as follows: the initial temperature was 70 C for 2 min, then raised to 180 C at 30 C/ min, to 200 C at 2 C/min, to 300 C at 30 C/min, and held for 10 min. Helium was used as carrier gas at a flow rate of 1 ml/min. A 1 ll sample was injected in splitless mode with solvent delay of 5 min.
Seawater Quality Standard (clean seawater) (Ma et al., 2005). Water and sediment were collected in the rainy season of 2002 from 38 sampling sites (Fig. 1). Surface water samples (0.5 m below surface) were collected using a stainless steel submersible pump and stored in amber glass bottles. The samples were filtered through prebaked glass fiber filters (Whatman GF/F, 0.7 lm effective pore size) immediately. The filtrates were acidified with 4 M HCl to pH < 1 to depress microbial degradation and stored at 4 C prior to extraction within two days. Surface sediment samplers (up to top 20 cm) were collected using a stainless steel grab sampler, scooped into aluminum jars and stored at 20 C until further analysis. 2.3. Analysis Analytical procedures were based on those described by Isobe et al. (2001). The filtrate was neutralized to pH 3–5 with sodium hydroxide just before extraction. After spiked with 4-tert-butylphenol as a surrogate standard, each neutralized filtrate sample (1 l) was concentrated onto a pre-conditioned solid-phase extraction cartridge (SEP-PAK tC18 plus environmental cartridge, Waters, Milford,MA), and then the analytes were eluted from
TIC
Abundance
Water NP
S.S
OP I.S
TIC
Abundance
Sediments NP S.S I.S OP
Time-->
6.00
7.00
8.00
9.00
10.00
11.00
12.00
13.00
Min.
Fig. 2. GC/MS chromatograms of target compounds in water and sediment samples. S.S and I.S represented 4-tert-butylphenol as the surrogate standard and 4-a-cumylphenol as the internal standard, respectively.
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All the target compounds could be well separated as shown in Fig. 2. NP consists of 11 isomer peaks due to various branched structures in the nonyl subsistent, and OP consists of a single peak due to one specific structure. NP was quantified by using the sum of all peaks area. Quantification was performed using the internal calibration method based on five-point calibration curve. The calibration curves for NP and OP showed high linearity (r2 P 0.995). The limits of quantification (LOQ) were calculated by a signal-to-noise of 10. Taking into account of the amount of sample extracted and volume of the extraction analyzed, LOQ for water samples were estimated 0.020 lg l1 for NP and 0.002 lg l1 for OP, while LOQ for sediment samples were estimated 10 lg kg1 for NP and 1 lg kg1 for OP. Reproducibility and efficiency of the analytical procedure for water samples were determined by six replicate analyses of redistilled de-ionized water spiked with APs standards. The average recovery of NP and OP was 92% and 80%, respectively, and their RSD was 6% and 10%, respectively. Reproducibility and efficiency of the analytical procedure for sediment samples were determined by six replicate analyses of pre-extracted sediments spiked with APs standards. The average recovery of NP and OP was 106% and 87%, respectively, and their RSD was 5% and 7%, respectively. For each of 8–10 water or sediment samples, a procedure blank, a spiked blank, and at least one replicate were included. To minimize procedural blanks, the entire analytical procedure was conducted carefully: All organic solvents were redistilled using a glass system before use; all the glassware was baked for 4–5 h at 450 C prior to use to remove organic contamination. The analytical results indicated that the procedure blanks contained no detectable amounts of APs and so no blank correction was necessary. In addition, in order to monitor procedural performance and matrix effects, 4-tertbutylphenol as surrogate standard was added to all the samples, and the recoveries were 81–110%. Sediment total organic carbon (TOC) content was measured with CHNS Vario E1IIIelement analyzer. The freeze-dried and homogenized sediments were first acidified with 10% (v/v) HCl overnight to remove carbonate, then dried at 60 C and analyzed for TOC (Hedges and Stren, 1984).
3. Results and discussion 3.1. APs in surface water The concentrations of NP and OP in surface water from the Pearl River Delta and adjacent northern South China Sea are listed in Table 1. NP concentrations varied between <0.020 and 0.628 lg l1, while OP concentrations varied between <0.002 and 0.068 lg l1.
Similar to those observed in other aquatic environment (Bennie et al., 1997; Isobe et al., 2001; Jin et al., 2004), OP concentrations were one order of magnitude lower than NP concentrations, reflecting less usage of OPEs in the Pearl River Delta. The spatial distribution of total APs in surface water is shown in Fig. 3a. Most of the water samples collected from the rivers contained detectable concentrations of APs. Substantially higher concentrations of APs were found at station ZR2 and ZR3 adjacent to the petroleum chemical industry zone of Guangzhou in the Zhujiang River. There was obviously a decrease of APs in going from the Zhujiang River to the Shiziyang River. The decrease could be explained by the fact that the Zhujiang River in highly industrialized and urbanized areas receives much more industrial and domestic wastewater effluents than the Shiziyang River in a less-urbanized and agricultural area. Compared with other rivers, the Xijiang River contained the lowest concentrations of APs, and did not vary greatly among the sampling sites. This might result from higher river flows which result in better dilution of local discharges and also because there were fewer actual industrial and domestic wastewater discharge point sources in this river system. All the water samples collected from the coast of Macao had detectable concentrations of APs in the range from 0.029 to 0.047 lg l1 with the highest value recorded at station MC1 inside the Macao Harbor. In the Lingding Bay only the water sample at station LB1 contained measurable concentrations of APs. From the estuary to its outer waters, concentrations of APs showed seaward decreasing trend. Bester et al. (2001) also reported the similar trend in NP concentrations from the River Elbe estuary to the German Bight of the North Sea. 3.2. APs in surface sediments As expected, NP and OP were found to be mainly concentrated in sediments from the Pearl River Delta and adjacent northern South China Sea (Table 1). All the sediment samples had detectable concentrations of NP and OP in the range from 59 to 7808 lg kg1 and from 1 to 93 lg kg1, respectively. NP concentrations were two orders of magnitude higher than OP concentrations in sediments, which was different from that observed in water. This result could be explained by their difference in octanol/water partition coefficients (log Kow). Log Kow of NP is 4.48 while that of OP is 4.12 (Ahel and Giger, 1993), indicating that NP should have a higher affinity for sediment than OP. The spatial distribution of total APs in surface sediments is shown in Fig. 3b. Measured concentrations of APs in riverine sediments varied significantly from 111 to 7901 lg kg1. Similar to water, sediments collected from the Zhujiang River showed the highest concentra-
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Table 1 Concentrations of APs in water and sediments from the Pearl River Delta and northern South China Sea Sampling site
Zhujiang River ZR1 ZR2 ZR3 ZR4 ZR5
Latitude/longitude
Sampling date
Tidal situation
Water (lg l1)
Sediments (lg kg1 dw)
NP
OP
NP
OP
TOC (%)
0.024 0.266 0.628 0.096 0.023
0.003 0.045 0.068 0.030 <0.002
4694 7808 5305 5936 2588
52 93 88 89 18
5.04 5.66 3.92 1.98 3.31
N2306.27 0 /E11314.27 0 N2306.48 0 /E11316.59 0 N2306.47 0 /E11318.59 0 N2302.57 0 /E11318.20 0 N2306.47 0 /E11318.59 0
2002-8-16 2002-8-16 2002-8-16 2002-8-16 2002-8-16
Ebbing Ebbing Ebbing Ebbing Ebbing
Shiziyang River SR1 N2302.45 0 /E11331.09 0 SR2 N2301.47 0 /E11331.09 0 SR3 N2257.37 0 /E11332.52 0 SR4 N2251.45 0 /E11333.23 0 SR5 N2253.27 0 /E11330.22 0 SR6 N2247.52 0 /E11335.46 0
2002-8-16 2002-8-16 2002-8-16 2002-8-16 2002-8-16 2002-8-16
Ebbing Ebbing Ebbing Ebbing Ebbing Flooding
<0.020 0.128 0.036 0.130 0.027 0.036
<0.002 0.003 <0.002 0.002 <0.002 <0.002
1361 508 1001 438 297 258
7 13 14 14 10 2
2.01 2.01 1.74 1.12 0.81 0.85
Xijiang River XR1 XR2 XR3 XR4 XR5 XR6 XR7 XR8
2002-9-5 2002-9-5 2002-9-5 2002-9-5 2002-9-6 2002-9-6 2002-9-6 2002-9-6
Ebbing Ebbing Ebbing Ebbing Ebbing Ebbing Ebbing Ebbing
0.031 0.026 <0.020 <0.020 <0.020 0.040 0.028 0.028
<0.002 <0.002 <0.002 <0.002 <0.002 <0.002 <0.002 <0.002
155 110 120 113 156 448 350 171
11 1 7 9 4 8 16 13
1.41 0.38 0.76 0.98 1.67 1.32 1.40 1.44
Dongjiang River DR1 N2302.57 0 /E11332.40 0 DR2 N2302.02 0 /E11331.22 0 DR3 N2258.21 0 /E11333.20 0 DR4 N2254.18 0 /E11335.16 0 DR5 N2257.02 0 /E11337.50 0 DR6 N2306.12 0 /E11352.03 0
2002-10-14 2002-10-11 2002-10-11 2002-10-12 2002-10-12 2002-10-13
Flooding Ebbing Ebbing Ebbing Ebbing Ebbing
naa na na na na na
na na na na na na
390 912 209 393 485 305
3 13 4 2 1 3
1.82 2.14 1.48 0.57 2.64 1.30
Macao Coast MC1 MC2 MC3 MC4
N2211.18 0 /E11331.46 0 N2209.30 0 /E11330.09 0 N2209.27 0 /E11332.21 0 N2210.48 0 /E11332.54 0
2002-9-7 2002-9-7 2002-9-7 2002-9-7
Ebbing Ebbing Ebbing Ebbing
0.037 0.039 0.029 0.031
0.009 <0.002 <0.002 <0.002
1208 305 293 302
28 16 7 19
1.47 1.04 1.38 0.73
Lingding Bay LB1 LB2 LB3 LB4 LB5
N2235.52 0 /E11342.87 0 N2228.93 0 /E11344.88 0 N2221.12 0 /E11346.64 0 N2212.28 0 /E11347.88 0 N2209.65 0 /E11350.61 0
2002-7-23 2002-7-23 2002-7-23 2002-7-23 2002-7-24
Ebbing Ebbing Ebbing Ebbing Ebbing
0.030 <0.020 <0.020 <0.020 <0.020
<0.002 <0.002 <0.002 <0.002 <0.002
60 399 571 535 571
1 8 5 15 18
0.06 0.94 0.92 0.94 1.02
South China Sea SS1 N2203.53 0 /E11351.84 0 SS2 N2157.51 0 /E11352.78 0 SS3 N2142.99 0 /E11355.94 0 SS4 N2124.81 0 /E11359.62 0
2002-7-25 2002-7-25 2002-7-25 2002-7-25
<0.020 <0.020 <0.020 <0.020
<0.002 <0.002 <0.002 <0.002
163 90 90 59
8 5 4 3
0.37 0.64 0.37 0.54
a
N2309.19 0 /E11249.01 0 N2306.50 0 /E11247.50 0 N2248.35 0 /E11300.37 0 N2204.01 0 /E11304.40 0 N2235.50 0 /E11309.56 0 N2231.46 0 /E11310.25 0 N2223.18 0 /E11315.59 0 N2209.57 0 /E11325.37 0
na: not analyzed.
tions of APs, followed by sediments from the Shiziyang River, and the Xijiang River contained the lowest concentrations of APs. Similar distribution trend were reported for chlorinated pesticides (CPs) and polychlorinated
biphenyls (PCBs) in riverine sediments from the Pearl River Delta (Kang et al., 2000; Mai et al., 2002). As indicated above, the Shiziyang River receives inflows from the Dongjiang River via four outlets. Around these outlets
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0.7
0.5 0.4 0.3 0.2
ZR1 ZR2 ZR3 ZR4 ZR5 SR1 SR2 SR3 SR4 SR5 SR6 XR1 XR2 XR3 XR4 XR5 XR6 XR7 XR8 MC1 MC2 MC3 MC4 LB1 LB2 LB3 LB4 LB5 SS1 SS2 SS3 SS4
0.1 0
Concentration (μg kg–1)
a
0.6
9000 8000
b
7000 6000 5000 4000 3000 2000 1000 0
ZR1 ZR2 ZR3 ZR4 ZR5 SR1 SR2 SR3 SR4 SR5 SR6 XR1 XR2 XR3 XR4 XR5 XR6 XR7 XR8 DR1 DR2 DR3 DR4 DR5 DR6 MC1 MC2 MC3 MC4 LB1 LB2 LB3 LB4 LB5 SS1 SS2 SS3 SS4
Concentration (μg l–1)
0.8
Sampling sites Fig. 3. Distribution of total APs in surface water (a) and sediments (b) from the Pearl River Delta and adjacent northern South China Sea.
except the Mayongkou Outlet, increasing trends of APs in sediments from the Dongjiang River to the Shiziyang River were observed (Fig. 4). The Northernkou Outlet is the largest one and discharges 2.4 · 1010 m3 of freshwater into the Shiziyang River, which receives 3.8 · 1010 m3 of freshwater from the Zhujiang River annually (GPAGTPA, 1994). Therefore, the dilution from the Dongjiang River is another reason which resulted in the decreases of APs in the Shiziyang River. Concentrations of APs in estuarine and marine sediments varied between 61 and 1236 lg kg1. Noticeably, the sediment sample collected from station MC1 inside Macao Harbor showed considerably higher concentration of APs than others. Previous studies also observed the highest levels of CPs, PCBs and PAHs inside the Macao Harbor sediments around the estuary (Kang et al., 2000; Mai et al., 2002). Macao Harbor is located in a long and narrow bay near Macao Island, so mixing between outer and inter water is limited, in addition, station MC1 is close to the downstream of a large sewage discharge channel in Macao, therefore caused high concentrations of APs at this location. It is interesting to note that sediments in the Lingding Bay showed relatively higher concentrations of APs than
those in the Xijiang River, although water collected from the Lingding Bay contained slightly lower amounts of APs than those from the Xijiang River. To some degree, the result might attribute to different hydrodynamic and sedimentary characteristics. In the Lingding Bay, gravitational circulation, tidal trapping, sediment resuspension and deposition process form turbidity maximum (Wai et al., 2004). The turbidity maximum has the function of enriching pollutants, and plays an important role in aggregation and sedimentation of fine cohesive sediment solids (Seylen and Hartin, 1990; Shen, 1995). The input annual suspended solids amounted to 3.4 · 107 tons in the Lingding Bay, of which 80% deposit in this Bay (Xu et al., 1993). On the other hand, a large proportion of APs with suspended solids leave the Xijiang River during the process of transportation. The input suspended solids amounted to 7.6 · 107 tons per year in the Xijiang River, representing approximately 86% of the total input amounts in the Pearl River. It is estimated that 20% of input suspended solids deposit in the Xijiang River (Xu et al., 1993). Unlike seaward decrease in water, concentrations of APs in sediments increased to a maximum, and then decrease offshore in the Lingding Bay and its outer waters.
APs (μgkg–1)
B. Chen et al. / Chemosphere 63 (2006) 652–661 1800 1600 1400 1200 1000 800 600 400 200 0
3.3. Comparison with other surface water and sediments
Dongjiang River Shiziyang River
SR1
SR3
DR2 SR2
DR1
DR4
SR4 5
DR3
Northernkou Outlet
Mayongkou Daoyunhaikou Southernkou Outlet Outlet Outlet
Fig. 4. Influence of outflows from the Dongjiang River on APs in the Shiziyang River.
The distribution might be associated with the sediment organic carbon, which is an important factor for sorption of APs (Johnson et al., 1998; Ferguson et al., 2001; Jonkers et al., 2003; Rice et al., 2003). The effect of the sediment organic carbon content on the levels of APs in the sediments was investigated (Table 1). Interestingly, the highest and lowest values of the sediment organic carbon and APs appeared simultaneously at Stations ZR2 and LB1, respectively. In general, sediments with high organic carbon contents showed high concentrations of APs; on the contrary, sediments with low organic carbon contents contained low concentrations of APs. Linear regression analyses showed that the concentrations of APs in the sediments were markedly correlated to the sediment organic carbon content with correlation coefficient of 0.832 (Fig. 5), indicating that the sediment organic carbon plays an important role in controlling the levels of APs in sediments. Sediments with high organic carbon content may favor accumulation of APs.
9000 8000
2
R = 0.723
APs (μg kg–1)
7000 6000 5000 4000 3000 2000 1000 0 0
1
2
3
4
5
659
6
TOC (%) Fig. 5. Correlation between total APs and TOC in sediments from the Pearl River Delta and adjacent northern South China Sea.
In other areas in China, there are no reports on the occurrence of APs in surface sediments, and data on levels of APs in surface water are also very limited. NP concentrations up to 6.85 lg l1 were recorded in river water from Chongqing (Shao et al., 2002), which were higher than values obtained in this study. Concentrations in river water from Tianjin were reported to vary from 0.106 to 0.553 lg l1 for NP and from 0.018 to 0.032 lg l1 for OP (Jin et al., 2004), which fell in the high values of concentration ranges in this study. The concentrations of APs found in this study were comparable to those reported from other industrialized and urbanized areas in the world. For example, NP concentrations up to 0.416 lg l1 in surface water and 13 700 lg kg1 in surface sediments were reported from Jamaica Bay, USA (Ferguson et al., 2001). Concentrations of NP in the range from 0.051 to 1.08 lg l1 in river water while from 30 to 13 000 lg kg1 in riverine and bay sediments were reported from the Tokyo metropolitan area, Japan (Isobe et al., 2001). A recent study reported NP concentrations up to 1.5331 lg l1 in surface water and 5054.1 lg kg1 in surface sediments from Shihwa Lake, Korea (Li et al., 2004).
4. Conclusions Endocrine disrupting APs were ubiquitous in the aquatic environment from the Pearl River Delta and adjacent northern South China Sea. The Zhujiang River in highly industrialized and urbanized area showed the highest concentrations of APs in both water and sediments. Significant decrease of APs concentrations in the Shiziyang River mainly attributed to dilution by inflows from the Dongjiang River in addition to less wastewater discharged into the Shiziyang River. Compared with the Lingding Bay, the Xijiang River contained concentrations of APs slightly higher in water but relatively lower in sediments, which might result from their different hydrodynamic and sedimentary characteristics. Concentrations of APs in water declined from the rivers to the estuary and further to the sea on the whole due to dilution. In the Lingding Bay and its outer waters, concentrations of APs in sediments increased to a maximum and then decreased seaward, which was consistent with the distribution trend of the sediment organic carbon. Linear regression analyses showed the concentrations of APs were markedly correlated with the sediment organic carbon contents, indicating that the sediment organic carbon is an important factor controlling the levels of APs in sediments. Sediments with high organic carbon content may favor accumulation of APs.
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Acknowledgements The financial support by the Knowledge Innovation Project of Chinese Academy of Sciences (No. ZKCX3SW-4429), National Basic Research Program of China (No. 2003CB415002) and the Natural Science Foundation of the PeopleÕ Republic of China (No. 40272129) is gratefully acknowledged. The authors thank Dr. Kedong Yin (South China Sea Institute of Oceanology) and Mr. Haiqiang Qian (Sanshui Hydrometric Gauging Station) for their help in collecting samples. The anonymous reviewers are also sincerely appreciated for their critical reviews that greatly improve this paper. References Ahel, M., Giger, W., 1993. Partitioning of alkylphenols and alkylphenol polyethoxylates between water and organic solvents. Chemosphere 26, 1471–1478. Basheer, C., Lee, H.K., Tan, K.S., 2004. Endocrine disrupting alkylphenols and bisphenol-A in coastal waters and supermarket seafood from Singapore. Marine Pollution Bulletin 48, 1145–1167. Bennie, D.T., Sullivan, C.A., Lee, H.B., Peart, T.E., Maguire, R.J., 1997. Occurrence of alkylphenols and alkylphenol mono- and diethoxylates in natural waters of the Laurentian Great Lakes basin and the upper St. Lawrence River. Science of the Total Environment 193, 263–275. Bester, K., Theobald, N., Schro¨der, H.Fr., 2001. Nonylphenols, nonylphenol-ethoxylates, linear alkyobenzenesulfonates (LAS) and bis (4-chlorophenyl)-sulfone in the German Bight of the North Sea. Chemosphere 45, 817–826. Blackburn, M.A., Kirby, S.J., Waldock, M.J., 1999. Concentrations of alkylphenol polyethoxylates entering UK estuaries. Marine Pollution Bulletin 38, 109–118. Bolz, U., Hagenmaier, H., Ko¨rner, W., 2001. Phenolic xenoestrogens in surface water, sediments, and sewage sludge from Baden-Wu¨rttemberg, South-west Germany. Environment Pollution 115, 291–301. Cheung, K.C., Poon, B.H.T., Lan, C.Y., Wong, M.H., 2003. Assessment of metal and nutrient concentrations in river water and sediment collected from the cities in the Pearl River Delta, South China. Chemosphere 52, 1431–1440. Comber, M.H.I., Williams, T.D., Stewart, K.M., 1993. The effect of nonylphenol on Daphnia magama. Water Research 27, 273–276. Ferguson, P.L., Iden, C.R., Brownwell, B.J., 2001. Distribution and fate of neutral alkylphenol ethoxylate metabolites in a sewage-impacted urban estuary. Environmental Science and Technology 35, 2428–2435. Fu, J.M., Mai, B.X., Sheng, G.Y., Zhang, G., Wang, X.M., Peng, P.A., Xiao, X.M., Ran, Y., Cheng, F.Z., Peng, X.Z., Tang, U.W., 2003. Persistent organic pollutants in environment of the Pearl River Delta, China: an overview. Chemosphere 52, 1411–1422. GPA-GTPA (Guangzhou Planning Association and Guangzhou Territy Planning Association), 1994. The Territory Resources of Guangzhou: Natural Resource. Guangzhou Press, Guangzhou, PR China (in Chinese).
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Sole, M., Lopez de Alda, M.J., Castillo, M., Porte, C., Ladegaard-Pedersen, K., Barcelo, D., 2000. Estrogenicity determination in sewage treatment plants and surface waters from the Catalonian area (NE Spain). Environmental Science and Technology 34, 5076–5083. Soto, A.M., Justicia, H., Wray, J.W., Sonnenschein, C., 1991. p-Nonylphenol: an estrogenic xenobiotic released from ÔmodifiedÕ polystyrene. Environmental Health Perspective 92, 167–173. Thiele, B., Gu¨nther, K., Schuwuger, M.J., 1997. Alkylphenol ethoxylates: trace analysis and environmental behavior. Chemical Reviews 97, 3248–3272. Tsuda, T., Takino, A., Kojima, M., Harada, H., Muraki, K., Tsuji, M., 2000. 4-Nonylphenols and 4-tert-octylphenol in water and fish from rivers flowing into Lake Biwa. Chemosphere 41, 757–762. Wai, O.W.H., Wang, Y.S., Li, Y.S., Li, X.D., 2004. The formation mechanisms of turbidity maximum in the Pearl River estuary, China. Marine Pollution Bulletin 48, 441– 448. Xu, J.L., Luo, Z.R., Wang, W.J., Dong, Z.Y., Lin, Z.W., Ye, M.Y., 1993. Development of the Rivers and Estuary as well as the Harbors in the Pearl River Delta. Ocean Press, Beijing, PR China (in Chinese). Ying, G.G., Williams, B., Kookana, R., 2002. Environmental fate of alkylphenols and alkylphenol ethoxylates—a review. Environmental International 28, 215–226. Zhang, G., Parker, A., House, A., Mai, B.X., Li, X.D., Kan, Y.H., Wang, Z.S., 2002. Sedimentary records of DDT and HCH in the Pearl River Delta, South China. Environmental Science and Technology 36, 3671–3677.
Distribution of alkylphenols in the Pearl River Delta and adjacent northern South China Sea, China Bing Chen a,b, Jing-Chun Duan a, Bi-xian Mai a,*, Xiao-Jun Luo a, Qing-Shu Yang a, Guo-Ying Sheng a, Jia-Mo Fu a a
State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, P.O. Box 1130, Guangzhou 510640, PR China b School of Environmental Science and Engineering, South China University of Technology, Guangzhou 510641, PR China Received 17 March 2005; received in revised form 26 July 2005; accepted 8 August 2005 Available online 10 October 2005
Abstract The occurrence of alkylphenols (APs) was investigated in surface water and sediments from the Pearl River Delta and adjacent northern South China Sea. Most of the water samples contained detectable amounts of APs, ranging up to 0.628 lg l1 for nonylphenol (NP) and 0.068 lg l1 for octylphenol (OP). APs were found in all of the sediment samples with concentrations ranging from 59 to 7808 lg kg1 for NP and from 1 to 93 lg kg1 for OP. The Zhujiang River showed the highest concentrations of APs in both water and sediments. Significant decrease of APs concentrations going from the Zhujiang River to the Shiziyang River was observed. The Xijiang River contained concentrations of APs slightly higher in water but relatively lower in sediments than the Lingding Bay, which might be attributed to their different hydrodynamic and sedimentary characteristics. There was a decreasing trend of APs in water from the rivers to the estuary and further to the sea on the whole. In the Lingding Bay and its outer waters, concentrations of APs in sediments increased to a maximum and then decrease seaward, which was consistent with the distribution trend of the sediment organic carbon contents. Linear regression analyses showed the concentrations of APs were markedly correlated with the sediment organic carbon contents, indicating that the sediment organic carbon is an important factor controlling the levels of APs in sediments. 2005 Elsevier Ltd. All rights reserved. Keywords: Alkylphenol; Water; Sediment; Pearl River Delta; South China Sea
1. Introduction Alkylphenol ethoxylates (APEs) are the worldÕs third largest group of surfactants with annual global produc* Corresponding author. Tel.: +86 20 85290146; fax: +86 20 85290706. E-mail addresses: [email protected] (B. Chen), nancymai@ gig.ac.cn (B.-x. Mai).
tion of 500 000 tons (Renner, 1997). They have been widely used in domestic detergents, pesticide formulations and industrial products, such as textiles, coatings, paints, lube oils, fuels, metals, plastics, pulp and paper. Nonylphenol ethoxylates (NPEs) and octylphenol ethoxylates (OPEs) are two of the most common surfactants in the marketplaces. NPEs account for about 80% of APEs, and OPEs account for most of the remainder (Renner, 1997). APEs enter the aquatic environment
0045-6535/$ - see front matter 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2005.08.004
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primarily via industrial and domestic wastewater with or without treatment due to their application characteristics. They can be biodegraded to lower ethoxymer and subsequently generate more persistent and lipophilic APs under anaerobic conditions (Thiele et al., 1997; Ying et al., 2002). There has been increasing concern over the presence of APs in the environment because of their toxicity (McLeese et al., 1981; Comber et al., 1993) and estrogenic activity (Soto et al., 1991; Jobling and Sumpter, 1993; Jobling et al., 1996). Since the route of wildlife and human exposure to these substances is mainly through water, the occurrence of APs has been investigated in a variety of rivers, lakes, estuaries and coastal waters around the world (e.g. Naylor et al., 1992; Bennie et al., 1997; Blackburn et al., 1999; Tsuda et al., 2000; Bester et al., 2001; Ferguson et al., 2001; Isobe et al., 2001; Jonkers et al., 2003; Rice et al., 2003; Basheer et al., 2004; Li et al., 2004). Reported levels in surface water ranged from <0.01 to 644 lg l1 for NP and from <0.005 to 0.18 lg l1 for OP (Bennie et al., 1997; Sole et al., 2000; Isobe et al., 2001). APs in water tend to be associated with particulate matter due to their hydrophobic nature. They are transported and finally deposited to the bottom sediments (Ferguson et al., 2001; Isobe et al., 2001). Relative to their short life in water, APs tend to accumulate over long periods of time in sediments (Liber et al., 1999; Shang et al., 1999). Levels of APs found in surface sediments were therefore substantially higher than the aqueous levels. Their concentrations ranged from <0.4 to 72 000 lg kg1 for NP and from <0.5 to 1800 lg kg1 for OP (Bennie et al., 1997; Bolz et al., 2001; Jonkers et al., 2003). Although distribution of APs in the aquatic environment has been documented in many countries, little information on APs is available in China where the production of APEs is over 50 000 tons annually (Huang, 1998). The Pearl River is the second largest river in China. It has a great number of tributaries and streams, forming a complicated watershed called the Pearl River Delta (Fig. 1). Approximately 3.3 · 1011 m3 of freshwater with suspended solids of 7.1 · 107 tons annually flow into the Pearl River Estuary via eight major outlets (Xu et al., 1993), and ultimately into the northern part of South China Sea. The Pearl River Delta covers an area of 41 700 km2 with a population of 41 million. As a result of rapid economic development in recent decades, the Pearl River Delta has become one of the most prosperous regions in China. However, the high growth of population, the rapid development of industry and agriculture, and the use of chemicals have caused serious pollution problems, leading to excessive release of pollutants into this region. It is estimated that industrial and domestic wastewater amounted to 3.0 · 109 m3 per year in the Pearl River Delta (Liu and Wu, 2004). Recent studies have revealed high levels of persistent organic
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pollutants and heavy metals in the river and estuarine environments (Cheung et al., 2003; Fu et al., 2003). To date, there has not been any report concerning APs contamination in this region. In this study, the occurrence of APs in water and sediments from the Pearl River Delta and adjacent northern South China Sea has been investigated. In addition, the distribution of APs in the aquatic environment combining rivers and estuary with sea as a whole has been discussed, which is important in assessing the impact of anthropogenic activities on the aquatic ecosystem and designing the management and conservation policies of this region.
2. Materials and methods 2.1. Chemicals 4-nonylphenol (mixture of compounds with branched side chain), 4-tert-octylphenol (93%), 4-tert-butylphenol (98%) and 4-a-cumylphenol (98%) were purchased from Tokyo Chemical Industry (Tokyo, Japan). Organic solvents were analytical grade and redistilled using a glass system. Neutral silica gel (80–100 mesh) was Soxhlet extracted with dichloromethane for 48 h, activated at 180 C for 12 h and then deactivated by adding 5% distilled water. 2.2. Study area and sample collection Three major tributaries of the Pearl River within the Pearl River Delta (Fig. 1) were selected for sampling. The first tributary comprises the Zhujiang River (ZR1– ZR5) and Shiziyang River (SR1–SR6). The Zhujiang River is situated in the northeast of the Pearl River Delta. It runs through the city of Guangzhou, the highly urbanized and densely populated provincial capital (7 million people) with numerous factories and industries, and are expected to be heavily contaminated with pollutants of industrial and domestic wastewater. The Shiziyang River receives inflows from the Zhujiang River and Dongjiang River, and empties into the Lingding Bay through the outlet of Humen. The second tributary is the Xijiang River (XR1–XR8), lying in the southwest of the Pearl River Delta. It runs across less developed areas and drains into the estuary through the Modaomen, Jitimen, and Hutiaomen outlets. The last tributary is the Dongjiang River (DR1–DR6), located in the northeast of the Pearl River Delta. It run through the city of Dongguan, where manufacturing industry is well developed, and flows into the Shiziyang River via Northernkou, Mayongkou, Daoyunhaikou, and Southernkou outlets. The Pearl River Estuary is one of the largest estuaries in the world (Fig. 1). The coast of Macao (MC1–MC4), located in the west side of the estuary, was selected for sampling. It is a known depositional zone for persistent
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Fig. 1. Map showing sampling sites in the Pearl River Delta and adjacent northern South China Sea. The numbers (1–8) indicate the eight major outlets to the South China Sea: (1) Humen; (2) Jiaomen; (3) Honqilimen; (4) Hengmen; (5) Modaomen; (6) Jitimen; (7) Hutiaomen; (8) Yamen. The numbers (9–12) indicate the four outlets to the Shiziyang River: (9) Northernkou; (10) Mayongkou; (11) Daoyunhaikou; (12) Southernkou. ZR, Zhujiang River; SR, Shiziyang River; XR, Xijiang River; DR, Dongjiang River; MC, Macao Coast; LB, Lingding Bay; SS, South China Sea.
organic pollutants in the Pearl River Delta. Relatively high levels of PCBs, DDTs and PAHs were found in surface and core sediments from Macao Harbor (Kang et al., 2000; Mai et al., 2002; Zhang et al., 2002; Mai et al., 2003). The main Pearl River Estuary, i.e. Lingding Bay (LB1–LB5), and its outer waters (SS1–SS4) were also selected for sampling. Lying in the northern part of the estuary, the Lingding Bay is funnel shaped and receives runoff from four outlets, i.e. Humen, Jiaomen,
Hongqilimen, and Hengmen. In addition, it is also greatly impacted by input from Macao and Hong Kong, one of the most populated regions in the world. Therefore, the Lingding Bay is considered an important pathway for pollutant transport into the South China Sea. The South China Sea is the largest marginal sea next to the western boundary of the Pacific Ocean. The water quality of the northern South China Sea, except the area near the estuary, achieved the second grade of National
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the cartridge with 20 ml methanol. The sediments were freeze-dried and homogenized. Each sample (10 g) was spiked with 4-tert-butylphenol as a surrogate standard and Soxhlet extracted with dichloromethane for 48 h. All the obtained extracts were concentrated and solvent-exchanged to hexane. The hexane extracts were purified through silica gel column (1 cm i.d. · 9 cm). The column was eluted with 20 ml dichloromethane/ hexane (4:6) and 40 ml dichloromethane/hexane (8:2). The later was collected as APs fraction and concentrated to 0.1 ml for water samples and 0.5 ml for sediment samplers under a gentle N2 stream. 4-a-Cumylphenol as internal standard was added to the sample prior to GC–MS analysis. APs were analyzed on HP5890 Series II gas chromatograph coupled with HP5972 mass selective detector in selected ion monitoring mode with DB-5 fused silica capillary column (30 m · 0.25 mm i.d., 0.25 lm film thickness). The injector and detector temperatures were set at 280 C and 300 C, respectively. The oven temperature program was as follows: the initial temperature was 70 C for 2 min, then raised to 180 C at 30 C/ min, to 200 C at 2 C/min, to 300 C at 30 C/min, and held for 10 min. Helium was used as carrier gas at a flow rate of 1 ml/min. A 1 ll sample was injected in splitless mode with solvent delay of 5 min.
Seawater Quality Standard (clean seawater) (Ma et al., 2005). Water and sediment were collected in the rainy season of 2002 from 38 sampling sites (Fig. 1). Surface water samples (0.5 m below surface) were collected using a stainless steel submersible pump and stored in amber glass bottles. The samples were filtered through prebaked glass fiber filters (Whatman GF/F, 0.7 lm effective pore size) immediately. The filtrates were acidified with 4 M HCl to pH < 1 to depress microbial degradation and stored at 4 C prior to extraction within two days. Surface sediment samplers (up to top 20 cm) were collected using a stainless steel grab sampler, scooped into aluminum jars and stored at 20 C until further analysis. 2.3. Analysis Analytical procedures were based on those described by Isobe et al. (2001). The filtrate was neutralized to pH 3–5 with sodium hydroxide just before extraction. After spiked with 4-tert-butylphenol as a surrogate standard, each neutralized filtrate sample (1 l) was concentrated onto a pre-conditioned solid-phase extraction cartridge (SEP-PAK tC18 plus environmental cartridge, Waters, Milford,MA), and then the analytes were eluted from
TIC
Abundance
Water NP
S.S
OP I.S
TIC
Abundance
Sediments NP S.S I.S OP
Time-->
6.00
7.00
8.00
9.00
10.00
11.00
12.00
13.00
Min.
Fig. 2. GC/MS chromatograms of target compounds in water and sediment samples. S.S and I.S represented 4-tert-butylphenol as the surrogate standard and 4-a-cumylphenol as the internal standard, respectively.
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All the target compounds could be well separated as shown in Fig. 2. NP consists of 11 isomer peaks due to various branched structures in the nonyl subsistent, and OP consists of a single peak due to one specific structure. NP was quantified by using the sum of all peaks area. Quantification was performed using the internal calibration method based on five-point calibration curve. The calibration curves for NP and OP showed high linearity (r2 P 0.995). The limits of quantification (LOQ) were calculated by a signal-to-noise of 10. Taking into account of the amount of sample extracted and volume of the extraction analyzed, LOQ for water samples were estimated 0.020 lg l1 for NP and 0.002 lg l1 for OP, while LOQ for sediment samples were estimated 10 lg kg1 for NP and 1 lg kg1 for OP. Reproducibility and efficiency of the analytical procedure for water samples were determined by six replicate analyses of redistilled de-ionized water spiked with APs standards. The average recovery of NP and OP was 92% and 80%, respectively, and their RSD was 6% and 10%, respectively. Reproducibility and efficiency of the analytical procedure for sediment samples were determined by six replicate analyses of pre-extracted sediments spiked with APs standards. The average recovery of NP and OP was 106% and 87%, respectively, and their RSD was 5% and 7%, respectively. For each of 8–10 water or sediment samples, a procedure blank, a spiked blank, and at least one replicate were included. To minimize procedural blanks, the entire analytical procedure was conducted carefully: All organic solvents were redistilled using a glass system before use; all the glassware was baked for 4–5 h at 450 C prior to use to remove organic contamination. The analytical results indicated that the procedure blanks contained no detectable amounts of APs and so no blank correction was necessary. In addition, in order to monitor procedural performance and matrix effects, 4-tertbutylphenol as surrogate standard was added to all the samples, and the recoveries were 81–110%. Sediment total organic carbon (TOC) content was measured with CHNS Vario E1IIIelement analyzer. The freeze-dried and homogenized sediments were first acidified with 10% (v/v) HCl overnight to remove carbonate, then dried at 60 C and analyzed for TOC (Hedges and Stren, 1984).
3. Results and discussion 3.1. APs in surface water The concentrations of NP and OP in surface water from the Pearl River Delta and adjacent northern South China Sea are listed in Table 1. NP concentrations varied between <0.020 and 0.628 lg l1, while OP concentrations varied between <0.002 and 0.068 lg l1.
Similar to those observed in other aquatic environment (Bennie et al., 1997; Isobe et al., 2001; Jin et al., 2004), OP concentrations were one order of magnitude lower than NP concentrations, reflecting less usage of OPEs in the Pearl River Delta. The spatial distribution of total APs in surface water is shown in Fig. 3a. Most of the water samples collected from the rivers contained detectable concentrations of APs. Substantially higher concentrations of APs were found at station ZR2 and ZR3 adjacent to the petroleum chemical industry zone of Guangzhou in the Zhujiang River. There was obviously a decrease of APs in going from the Zhujiang River to the Shiziyang River. The decrease could be explained by the fact that the Zhujiang River in highly industrialized and urbanized areas receives much more industrial and domestic wastewater effluents than the Shiziyang River in a less-urbanized and agricultural area. Compared with other rivers, the Xijiang River contained the lowest concentrations of APs, and did not vary greatly among the sampling sites. This might result from higher river flows which result in better dilution of local discharges and also because there were fewer actual industrial and domestic wastewater discharge point sources in this river system. All the water samples collected from the coast of Macao had detectable concentrations of APs in the range from 0.029 to 0.047 lg l1 with the highest value recorded at station MC1 inside the Macao Harbor. In the Lingding Bay only the water sample at station LB1 contained measurable concentrations of APs. From the estuary to its outer waters, concentrations of APs showed seaward decreasing trend. Bester et al. (2001) also reported the similar trend in NP concentrations from the River Elbe estuary to the German Bight of the North Sea. 3.2. APs in surface sediments As expected, NP and OP were found to be mainly concentrated in sediments from the Pearl River Delta and adjacent northern South China Sea (Table 1). All the sediment samples had detectable concentrations of NP and OP in the range from 59 to 7808 lg kg1 and from 1 to 93 lg kg1, respectively. NP concentrations were two orders of magnitude higher than OP concentrations in sediments, which was different from that observed in water. This result could be explained by their difference in octanol/water partition coefficients (log Kow). Log Kow of NP is 4.48 while that of OP is 4.12 (Ahel and Giger, 1993), indicating that NP should have a higher affinity for sediment than OP. The spatial distribution of total APs in surface sediments is shown in Fig. 3b. Measured concentrations of APs in riverine sediments varied significantly from 111 to 7901 lg kg1. Similar to water, sediments collected from the Zhujiang River showed the highest concentra-
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Table 1 Concentrations of APs in water and sediments from the Pearl River Delta and northern South China Sea Sampling site
Zhujiang River ZR1 ZR2 ZR3 ZR4 ZR5
Latitude/longitude
Sampling date
Tidal situation
Water (lg l1)
Sediments (lg kg1 dw)
NP
OP
NP
OP
TOC (%)
0.024 0.266 0.628 0.096 0.023
0.003 0.045 0.068 0.030 <0.002
4694 7808 5305 5936 2588
52 93 88 89 18
5.04 5.66 3.92 1.98 3.31
N2306.27 0 /E11314.27 0 N2306.48 0 /E11316.59 0 N2306.47 0 /E11318.59 0 N2302.57 0 /E11318.20 0 N2306.47 0 /E11318.59 0
2002-8-16 2002-8-16 2002-8-16 2002-8-16 2002-8-16
Ebbing Ebbing Ebbing Ebbing Ebbing
Shiziyang River SR1 N2302.45 0 /E11331.09 0 SR2 N2301.47 0 /E11331.09 0 SR3 N2257.37 0 /E11332.52 0 SR4 N2251.45 0 /E11333.23 0 SR5 N2253.27 0 /E11330.22 0 SR6 N2247.52 0 /E11335.46 0
2002-8-16 2002-8-16 2002-8-16 2002-8-16 2002-8-16 2002-8-16
Ebbing Ebbing Ebbing Ebbing Ebbing Flooding
<0.020 0.128 0.036 0.130 0.027 0.036
<0.002 0.003 <0.002 0.002 <0.002 <0.002
1361 508 1001 438 297 258
7 13 14 14 10 2
2.01 2.01 1.74 1.12 0.81 0.85
Xijiang River XR1 XR2 XR3 XR4 XR5 XR6 XR7 XR8
2002-9-5 2002-9-5 2002-9-5 2002-9-5 2002-9-6 2002-9-6 2002-9-6 2002-9-6
Ebbing Ebbing Ebbing Ebbing Ebbing Ebbing Ebbing Ebbing
0.031 0.026 <0.020 <0.020 <0.020 0.040 0.028 0.028
<0.002 <0.002 <0.002 <0.002 <0.002 <0.002 <0.002 <0.002
155 110 120 113 156 448 350 171
11 1 7 9 4 8 16 13
1.41 0.38 0.76 0.98 1.67 1.32 1.40 1.44
Dongjiang River DR1 N2302.57 0 /E11332.40 0 DR2 N2302.02 0 /E11331.22 0 DR3 N2258.21 0 /E11333.20 0 DR4 N2254.18 0 /E11335.16 0 DR5 N2257.02 0 /E11337.50 0 DR6 N2306.12 0 /E11352.03 0
2002-10-14 2002-10-11 2002-10-11 2002-10-12 2002-10-12 2002-10-13
Flooding Ebbing Ebbing Ebbing Ebbing Ebbing
naa na na na na na
na na na na na na
390 912 209 393 485 305
3 13 4 2 1 3
1.82 2.14 1.48 0.57 2.64 1.30
Macao Coast MC1 MC2 MC3 MC4
N2211.18 0 /E11331.46 0 N2209.30 0 /E11330.09 0 N2209.27 0 /E11332.21 0 N2210.48 0 /E11332.54 0
2002-9-7 2002-9-7 2002-9-7 2002-9-7
Ebbing Ebbing Ebbing Ebbing
0.037 0.039 0.029 0.031
0.009 <0.002 <0.002 <0.002
1208 305 293 302
28 16 7 19
1.47 1.04 1.38 0.73
Lingding Bay LB1 LB2 LB3 LB4 LB5
N2235.52 0 /E11342.87 0 N2228.93 0 /E11344.88 0 N2221.12 0 /E11346.64 0 N2212.28 0 /E11347.88 0 N2209.65 0 /E11350.61 0
2002-7-23 2002-7-23 2002-7-23 2002-7-23 2002-7-24
Ebbing Ebbing Ebbing Ebbing Ebbing
0.030 <0.020 <0.020 <0.020 <0.020
<0.002 <0.002 <0.002 <0.002 <0.002
60 399 571 535 571
1 8 5 15 18
0.06 0.94 0.92 0.94 1.02
South China Sea SS1 N2203.53 0 /E11351.84 0 SS2 N2157.51 0 /E11352.78 0 SS3 N2142.99 0 /E11355.94 0 SS4 N2124.81 0 /E11359.62 0
2002-7-25 2002-7-25 2002-7-25 2002-7-25
<0.020 <0.020 <0.020 <0.020
<0.002 <0.002 <0.002 <0.002
163 90 90 59
8 5 4 3
0.37 0.64 0.37 0.54
a
N2309.19 0 /E11249.01 0 N2306.50 0 /E11247.50 0 N2248.35 0 /E11300.37 0 N2204.01 0 /E11304.40 0 N2235.50 0 /E11309.56 0 N2231.46 0 /E11310.25 0 N2223.18 0 /E11315.59 0 N2209.57 0 /E11325.37 0
na: not analyzed.
tions of APs, followed by sediments from the Shiziyang River, and the Xijiang River contained the lowest concentrations of APs. Similar distribution trend were reported for chlorinated pesticides (CPs) and polychlorinated
biphenyls (PCBs) in riverine sediments from the Pearl River Delta (Kang et al., 2000; Mai et al., 2002). As indicated above, the Shiziyang River receives inflows from the Dongjiang River via four outlets. Around these outlets
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0.7
0.5 0.4 0.3 0.2
ZR1 ZR2 ZR3 ZR4 ZR5 SR1 SR2 SR3 SR4 SR5 SR6 XR1 XR2 XR3 XR4 XR5 XR6 XR7 XR8 MC1 MC2 MC3 MC4 LB1 LB2 LB3 LB4 LB5 SS1 SS2 SS3 SS4
0.1 0
Concentration (μg kg–1)
a
0.6
9000 8000
b
7000 6000 5000 4000 3000 2000 1000 0
ZR1 ZR2 ZR3 ZR4 ZR5 SR1 SR2 SR3 SR4 SR5 SR6 XR1 XR2 XR3 XR4 XR5 XR6 XR7 XR8 DR1 DR2 DR3 DR4 DR5 DR6 MC1 MC2 MC3 MC4 LB1 LB2 LB3 LB4 LB5 SS1 SS2 SS3 SS4
Concentration (μg l–1)
0.8
Sampling sites Fig. 3. Distribution of total APs in surface water (a) and sediments (b) from the Pearl River Delta and adjacent northern South China Sea.
except the Mayongkou Outlet, increasing trends of APs in sediments from the Dongjiang River to the Shiziyang River were observed (Fig. 4). The Northernkou Outlet is the largest one and discharges 2.4 · 1010 m3 of freshwater into the Shiziyang River, which receives 3.8 · 1010 m3 of freshwater from the Zhujiang River annually (GPAGTPA, 1994). Therefore, the dilution from the Dongjiang River is another reason which resulted in the decreases of APs in the Shiziyang River. Concentrations of APs in estuarine and marine sediments varied between 61 and 1236 lg kg1. Noticeably, the sediment sample collected from station MC1 inside Macao Harbor showed considerably higher concentration of APs than others. Previous studies also observed the highest levels of CPs, PCBs and PAHs inside the Macao Harbor sediments around the estuary (Kang et al., 2000; Mai et al., 2002). Macao Harbor is located in a long and narrow bay near Macao Island, so mixing between outer and inter water is limited, in addition, station MC1 is close to the downstream of a large sewage discharge channel in Macao, therefore caused high concentrations of APs at this location. It is interesting to note that sediments in the Lingding Bay showed relatively higher concentrations of APs than
those in the Xijiang River, although water collected from the Lingding Bay contained slightly lower amounts of APs than those from the Xijiang River. To some degree, the result might attribute to different hydrodynamic and sedimentary characteristics. In the Lingding Bay, gravitational circulation, tidal trapping, sediment resuspension and deposition process form turbidity maximum (Wai et al., 2004). The turbidity maximum has the function of enriching pollutants, and plays an important role in aggregation and sedimentation of fine cohesive sediment solids (Seylen and Hartin, 1990; Shen, 1995). The input annual suspended solids amounted to 3.4 · 107 tons in the Lingding Bay, of which 80% deposit in this Bay (Xu et al., 1993). On the other hand, a large proportion of APs with suspended solids leave the Xijiang River during the process of transportation. The input suspended solids amounted to 7.6 · 107 tons per year in the Xijiang River, representing approximately 86% of the total input amounts in the Pearl River. It is estimated that 20% of input suspended solids deposit in the Xijiang River (Xu et al., 1993). Unlike seaward decrease in water, concentrations of APs in sediments increased to a maximum, and then decrease offshore in the Lingding Bay and its outer waters.
APs (μgkg–1)
B. Chen et al. / Chemosphere 63 (2006) 652–661 1800 1600 1400 1200 1000 800 600 400 200 0
3.3. Comparison with other surface water and sediments
Dongjiang River Shiziyang River
SR1
SR3
DR2 SR2
DR1
DR4
SR4 5
DR3
Northernkou Outlet
Mayongkou Daoyunhaikou Southernkou Outlet Outlet Outlet
Fig. 4. Influence of outflows from the Dongjiang River on APs in the Shiziyang River.
The distribution might be associated with the sediment organic carbon, which is an important factor for sorption of APs (Johnson et al., 1998; Ferguson et al., 2001; Jonkers et al., 2003; Rice et al., 2003). The effect of the sediment organic carbon content on the levels of APs in the sediments was investigated (Table 1). Interestingly, the highest and lowest values of the sediment organic carbon and APs appeared simultaneously at Stations ZR2 and LB1, respectively. In general, sediments with high organic carbon contents showed high concentrations of APs; on the contrary, sediments with low organic carbon contents contained low concentrations of APs. Linear regression analyses showed that the concentrations of APs in the sediments were markedly correlated to the sediment organic carbon content with correlation coefficient of 0.832 (Fig. 5), indicating that the sediment organic carbon plays an important role in controlling the levels of APs in sediments. Sediments with high organic carbon content may favor accumulation of APs.
9000 8000
2
R = 0.723
APs (μg kg–1)
7000 6000 5000 4000 3000 2000 1000 0 0
1
2
3
4
5
659
6
TOC (%) Fig. 5. Correlation between total APs and TOC in sediments from the Pearl River Delta and adjacent northern South China Sea.
In other areas in China, there are no reports on the occurrence of APs in surface sediments, and data on levels of APs in surface water are also very limited. NP concentrations up to 6.85 lg l1 were recorded in river water from Chongqing (Shao et al., 2002), which were higher than values obtained in this study. Concentrations in river water from Tianjin were reported to vary from 0.106 to 0.553 lg l1 for NP and from 0.018 to 0.032 lg l1 for OP (Jin et al., 2004), which fell in the high values of concentration ranges in this study. The concentrations of APs found in this study were comparable to those reported from other industrialized and urbanized areas in the world. For example, NP concentrations up to 0.416 lg l1 in surface water and 13 700 lg kg1 in surface sediments were reported from Jamaica Bay, USA (Ferguson et al., 2001). Concentrations of NP in the range from 0.051 to 1.08 lg l1 in river water while from 30 to 13 000 lg kg1 in riverine and bay sediments were reported from the Tokyo metropolitan area, Japan (Isobe et al., 2001). A recent study reported NP concentrations up to 1.5331 lg l1 in surface water and 5054.1 lg kg1 in surface sediments from Shihwa Lake, Korea (Li et al., 2004).
4. Conclusions Endocrine disrupting APs were ubiquitous in the aquatic environment from the Pearl River Delta and adjacent northern South China Sea. The Zhujiang River in highly industrialized and urbanized area showed the highest concentrations of APs in both water and sediments. Significant decrease of APs concentrations in the Shiziyang River mainly attributed to dilution by inflows from the Dongjiang River in addition to less wastewater discharged into the Shiziyang River. Compared with the Lingding Bay, the Xijiang River contained concentrations of APs slightly higher in water but relatively lower in sediments, which might result from their different hydrodynamic and sedimentary characteristics. Concentrations of APs in water declined from the rivers to the estuary and further to the sea on the whole due to dilution. In the Lingding Bay and its outer waters, concentrations of APs in sediments increased to a maximum and then decreased seaward, which was consistent with the distribution trend of the sediment organic carbon. Linear regression analyses showed the concentrations of APs were markedly correlated with the sediment organic carbon contents, indicating that the sediment organic carbon is an important factor controlling the levels of APs in sediments. Sediments with high organic carbon content may favor accumulation of APs.
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