Temporal Trends of Polybrominated Diphenyl Ethers ...

4 downloads 0 Views 628KB Size Report
of Polybrominated. Diphenyl Ethers and. Hexabromocyclododecanes in Marine Mammals with. Special Reference to Hong. Kong, South China. Ling Jin, James ...

Trends 22 Temporal of Polybrominated Diphenyl Ethers and Hexabromocyclododecanes in Marine Mammals with Special Reference to Hong Kong, South China Ling Jin, James C.W. Lam, Margaret B. Murphy, and Paul K.S. Lam* Contents 22.1 Introduction........................................................................................................................... 498 22.1.1 An Overview of Brominated Flame Retardants........................................................ 498 22.1.2 Use and Production of PBDEs and HBCDs.............................................................. 498 22.1.3 PBDEs in the Environment........................................................................................ 499 22.1.4 HBCDs in the Environment...................................................................................... 501 22.1.5 Study Objectives........................................................................................................ 502 22.2 Materials and Methods.......................................................................................................... 503 22.2.1 Sample Collection and Preparation........................................................................... 503 22.2.2 Chemical Analysis..................................................................................................... 503 22.2.3 QA/QC....................................................................................................................... 503 22.2.4 Statistical Analysis....................................................................................................504 22.3 Results and Discussion..........................................................................................................504 22.3.1 Concentrations of PBDEs and HBCDs in Cetaceans................................................504 22.3.2 Temporal Trend of PBDEs and HBCDs in Cetaceans..............................................504 22.3.3 Global Trends of PBDE and HBCD Concentrations in Marine Mammals............... 505 22.3.3.1 Japan........................................................................................................... 505 22.3.3.2 North America............................................................................................506 22.3.3.3 Arctic..........................................................................................................506 22.3.3.4 Europe.........................................................................................................507

* E-mail: [email protected] (Chapter corresponding author).

497

© 2012 by Taylor & Francis Group, LLC

498

Global Contamination Trends of Persistent Organic Chemicals

22.4 Conclusions............................................................................................................................ 507 Acknowledgments........................................................................................................................... 508 References....................................................................................................................................... 508

22.1  Introduction 22.1.1  An Overview of Brominated Flame Retardants Owing to advances in polymer science over the past 50 years, a large number of polymers with differing properties have been developed for daily applications ranging from clothing and furniture to electronics, vehicles, and computers. However, most of these polymers are petroleum-based and hence are flammable. In order to reduce fire risks and meet fire safety regulations, certain chemicals collectively known as flame retardants are applied to combustible materials such as plastics, wood, paper, and textiles [1]. Currently, there are more than 175 compounds or groups of compounds with known flame-retarding properties, which are generally divided into four classes: inorganic, halogenated organic, nitrogen-containing, and phosphorus-containing compounds [2]. Among the halogenated flame retardants, brominated compounds comprise the largest market share because of their lower decomposition temperatures, higher performance efficiency, and low cost [2,3]. Thus, brominated flame retardants (BFRs) have been extensively used to improve the fire resistance of materials such as plastics, textiles, furnishing foam, and electronic circuit boards [4]. Based on their use in the chemical industry, BFRs can be classified as either reactive or additive. Reactive BFRs such as the tetrabromobisphenol A (TBBPA) are covalently bound to the polymer matrix. Compared to their reactive counterparts, additive BFRs are not chemically bound to the product and therefore tend to migrate out of the product much more easily and are thus more likely to be released into the environment. Examples of additive BFRs include polybrominated diphenyl ethers (PBDEs), polybrominated biphenyls (PBBs), and hexabromocyclododecanes (HBCDs). Production of PBBs in the United States was phased out in the 1970s after a farm product contamination incident in Michigan [5]. In turn, production of PBDEs has increased, peaking in the mid-1990s [6].

22.1.2  Use and Production of PBDEs and HBCDs There are three PBDE technical formulations used in industry: penta-BDE, octa-BDE, and decaBDE. They are named according to the predominant homolog groups in the mixture (Figure 22.1). Penta-BDE is mostly used in polyurethane foam (mattresses, furniture, pillows) and in adhesives, while octa-BDE is mainly used in rigid plastics (acrylonitrile butadiene styrene, ABS) such as in computer casings and computer monitors [3]. The deca-BDE formulation is used in plastics such as high-impact polystyrene (HIPS) in electrical and electronic equipment, such as the back covers of televisions, also in rubber coating for wiring, as well as in textile back-coating in furniture [2]. Increasing concerns have been raised regarding these flame-retardant products due to their persistence, bioaccumulative characteristics, and potential adverse effects [1]. The penta-BDE and octa-BDE commercial mixtures have been listed as persistent organic pollutants (POPs) under the Br Br Br

O

Brx

Brx PBDEs

Figure 22.1  General structures of PBDEs and HBCDs.

© 2012 by Taylor & Francis Group, LLC

Br

Br Br HBCDs

499

Temporal Trends of Polybrominated Diphenyl Ethers

Table 22.1 Estimated Annual Worldwide Market Demand of PBDEs and HBCDs in 2001 by Region, and Total Estimated Demand in 2002 and 2003 (Metric Tons) Americas Europe Asia Total (2001) Total (2002) Total (2003) a

Penta-BDE

Octa-BDE

Deca-BDE

HBCDs

7100 150 150 7500 — —

1500 610 1500 3790 — —

24,500 7,600 23,000 56,100 65,700 56,400

2,800 9,500 3,900 16,700 21,400 22,000

Source: Adapted from de Wit, C.A. et al., Sci. Total Environ., 408, 2885, 2010. With permission. a Americas include North and South America in which the United States is the major consumer country in the former continent.

Stockholm Convention [7]. There are currently no restrictions on the production or use of decaBDE in most countries. However, according to the EU Directive on the restriction of use of certain hazardous substances in electrical and electronic equipment (RoHS Directive), manufacturers must substitute other compounds for PBBs and PBDEs in new equipment. Deca-BDE was exempted from the directive, but this exemption was later overturned and phase-out of deca-BDE was initiated in European countries in 2008 [8]. Stringent regulations on PBDE use have led to the emergence of substitute chemicals, and one group that is likely to take the place of PBDEs is HBCDs. As a nonaromatic, brominated cyclic alkane (Figure 22.1), HBCDs are not chemically bound to products and are primarily applied in thermoplastic polymers used in the manufacture of styrene resins [2]. HBCDs have also been used to a lesser extent in textile coatings, cable, latex binders, and unsaturated polyesters [3]. The commercial HBCD product is mainly composed of three diastereoisomers: α-(10%–13%), β-(1%–12%), and γ-HBCD (75%–89%) [9]. However, the isomeric profile varies depending on the product application. Temperatures above 160°C can subject HBCDs to thermal rearrangement, resulting in a specific mixture of stereoisomers (78% α-HBCD, 13% β-HBCD, and 9% γ-HBCD) [9]. Such conditions can occur during the production or processing of HBCD-containing materials such as extruded polystyrene, and therefore, the relative abundance of the various HBCD stereoisomers may differ from that of the technical HBCD mixtures. To date, there is no control on the production and use of HBCDs anywhere in the world. Table 22.1 summarizes the estimated annual market demand for PBDEs and HBCDs for the years 2001–2003.

22.1.3  PBDEs in the Environment The major point sources of PBDEs to the environment are industrial plants manufacturing the technical mixtures as well as facilities incorporating PBDEs into polymers [10]. Electronic waste (e-waste) recycling facilities have recently been highlighted as an important point source of PBDEs [11,12]. In addition, wear and tear of products containing PBDEs constitutes a diffuse, non-point source of these chemicals. Fragments from the disintegration of polyurethane foam have been suggested to be a mechanism by which PBDEs diffuse into the atmosphere [13]. PBDEs tend to be stable and persistent in the environment. Due to their high binding affinity to organic matter, PBDEs are often associated with soils and sediments [3]. In addition, particulates in air and water are important pathways for the transport of these contaminants on local, regional, and global scales. Lower-brominated PBDEs are generally more volatile, water soluble and bioaccumulate more than higher-brominated PBDEs [14]. Previous research found that higher-brominated PBDE

© 2012 by Taylor & Francis Group, LLC

500

Global Contamination Trends of Persistent Organic Chemicals

congeners can be transformed or broken down into lower-brominated congeners by biological [15,16] or environmental processes, such as photodegradation by sunlight [17]. These processes may increase the risks of PBDE exposure in the environment and organisms, as the levels of lower-brominated PBDE congeners increase, but little information about these impacts is available. However, the occurrence of the highest brominated congener, BDE209, in waterbird eggs [18], seal blubber [19], and human blood [20] indicated that this congener can be taken up by biota and is likely to bioaccumulate. As ubiquitous environmental contaminants, PBDEs have been reported in a variety of environmental matrices. At present, PBDE contamination is recognized as a matter of global concern since they have reached remote areas such as the deep ocean and the polar regions [21,22] and demonstrated considerable accumulation in humans [23]. A wide range of PBDE levels have been measured in human adipose tissue, from 17 to 462 ng/g lw in the United States [24] and 2.2 to 11.7 ng/g lw in Belgium [25]. Such high variation in PBDE levels among humans might reflect the higher market demand for PBDEs in North America, notwithstanding other environmental factors and habits. Significant amounts of PBDEs have also been detected in lipid-rich tissues of the top predators such as bird eggs [26,27] and marine mammal blubber [28,29]. Significant increases in PBDE levels in human serum have been observed over the past two decades [30]. A temporal increase in PBDE levels has also been detected in sediment cores [31] and in biological samples such as fish [32], bird eggs [33], and marine mammals [34]. The increasing environmental occurrence of PBDEs has heightened research interest in studying PBDE levels in different parts of the world and their effects on different organisms. South China, including Hong Kong and its neighboring Pearl River Delta (PRD) region, is one of the most rapidly growing regions in the world. The PRD has developed into a globally leading production site for computers and electronic products. To date, there is no legislation restricting the use of BFRs in China, and thus it is likely that large quantities of PBDEs are being used and released in manufacturing processes in the region. These releases, together with the e-waste dismantling and recycling industries located in Guangdong Province in southern China, have contributed to the high levels of PBDEs that have been detected in environmental matrices in the PRD. Previous studies reported total PBDE concentrations in riverine runoff and sediment samples from the PRD of 0.34–68 ng/L [35] and 0.44–7435 ng/g dry wt [36], respectively. These levels fall within the high end of the range of global PBDE concentrations. Situated at the mouth of the Pearl River, Hong Kong also receives high volumes of these persistent contaminants into its marine habitats from a variety of anthropogenic sources [37]. In addition, local population growth and industrial development have put considerable pressure on the marine environment of Hong Kong. A recent large-scale monitoring study revealed high levels of PBDEs in sediments and greenlipped mussels (Perna viridis) from Hong Kong marine waters, with concentrations ranging from 1.7 to 53.6 ng/g dry wt. and 27.0 to 83.7 ng/g dry wt., respectively; these levels are among the highest in the world [38]. These PBDEs may originate from the disposal of e-waste in southern China, as well as untreated local discharge [39]. Studies have also found that PBDEs accumulated in both edible freshwater and marine fish samples collected in Hong Kong and from nearby waters [40,41]. Analysis of PBDEs in top predators such as in waterbird eggs [26] and cetaceans [42] inhabiting Hong Kong have also been carried out. The presence of PBDEs in these many species confirms their potential for dispersal, biological uptake, and bioaccumulation in many different trophic compartments in the environment [43]. The occurrence of high levels of PBDEs may result in adverse health effects in wildlife as shown in laboratory studies. Neurodevelopmental toxicity has been linked to tetra- and penta-BDE congener exposure in rats [44]. A single oral dose of tetra- or penta-BDE on day 10 following birth permanently impaired spontaneous motor behavior, affected learning and memory, and had permanent behavioral effects in mice [45]. In a study in killifish (Fundulus heteroclitus) [46], behavioral test results suggested that embryonic exposure to DE-71 could alter activity level, fright response, predation rates, and learning ability in subsequent life stages. PBDEs are also suspected to be immunotoxins. In nestling American kestrels (Falco sparverius) exposed to environmentally relevant

© 2012 by Taylor & Francis Group, LLC

Temporal Trends of Polybrominated Diphenyl Ethers

501

PBDEs (18.7 μg ΣPBDEs/egg), greater T-cell-mediated phytohemagglutinin response and less of an antibody-mediated response were observed [47]. Structural changes in the spleen (fewer germinal centers), bursa (reduced apoptosis) and thymus (increased macrophages), and negative associations between the spleen somatic index and total PBDEs, and the bursa somatic indices and BDE-47 were also found. Several studies also reported that PBDEs can disrupt thyroid function and exposure to commercial PBDE mixtures (penta-, octa-, and deca-BDEs) can result in thyroid hormone imbalance [48,49]. In an in vitro study conducted in cells from rainbow trout, chicken, rat, and human, strong EROD induction occurred after exposure to BDE 77, 100, 119, and 126, although the maximum EROD activity was less than that induced by dioxins [50]. Furthermore, the effects of PBDEs and PCBs on EROD were found to be synergistic, supporting the idea that both groups of chemicals may act through the same biological mechanism [51].

22.1.4  HBCDs in the Environment Like PBDEs, HBCDs can enter the environment via a number of pathways, such as emission during production of BFRs, by migrating out of consumer products, or following disposal. The environmental occurrence of HBCDs was first documented in fish and sediment samples from the river Viskan in Sweden [52] and, since then, their presence has been reported in a variety of biotic and abiotic matrices at levels comparable to those of PBDEs [9]. Long-range transport of HBCDs has been proposed due to the detection of these chemicals in biota from Eastern Greenland and Svalbard [53,54]. However, the significance of the long-range atmospheric transport of HBCDs in relation to other sources and transport routes remains to be further established. Concentrations of HBCDs determined in matrices such as air [55], fish [56], dolphins [57], and sea lions [58] from the North American environment are generally lower than those in similar samples from Europe [8,59,60] and Asia [60–62]. In general, the different environmental residue levels appear to reflect different continental market demands. Although the technical HBCD mixture primarily consists of γ-HBCD, this composition profile is usually not reflected in environmental media [2]. Particularly, α-HBCD is consistently found to dominate the isomer pattern of HBCDs in biotic samples [63,64], providing clear evidence of enantioselective accumulation/biotransformation of the industrial combinations [65,66]. Increasing temporal trends of HBCDs have been observed in biological samples such as bird eggs [67] and marine mammals [58]. The elevated environmental levels of HBCDs have attracted growing interest and effort in investigating their occurrences and effects on living organisms. There is a growing body of literature about the toxic effects of HBCDs. An increase in hepatosomatic index (HSI) and inhibition of EROD activity were found in juvenile rainbow trout (Oncorhynchus mykiss) intraperitoneally exposed to 50 and 500 mg/kg HBCD for 28 days [68]. Long-term exposure of European flounder (Platichthys flesus) to HBCD did not affect their general health nor several toxicity parameters (e.g., behavior, survival, growth rate, relative liver, and gonad weight), and neither was hepatic EROD activity induced [69]. In contrast, long-term exposure to HBCD-induced EROD activity in rare minnow (Gobiocypris rarus) resulting in oxidative damage to lipids, proteins, and DNA and decreased antioxidant capacity due to excess reactive oxygen species (ROS) formation [70]. Disruption of the thyroid axis was observed in juvenile rainbow trout exposed to HBCD diastereoisomers with the most evident effects in the γ-HBCDexposed group, as indicated by a decreased level of circulating free total thyroid hormone FT4, an increased level of free total thyroid hormone FT3, and an increase in thyroid epithelial cell height [71]. In addition, changes in spontaneous behavior, learning, and memory defects, and a reduced number of nicotinic receptors were demonstrated in mice after neonatal exposure to HBCDs [72]. The observations of HBCD-induced neurobehavioral alterations were supported by the molecular evidence that HBCDs block the uptake of dopamine into rat brain synaptosomes in vitro [73]. Developmental toxicity and apoptosis were also observed in zebrafish (Danio rerio) after embryonic exposure to HBCDs [74].

© 2012 by Taylor & Francis Group, LLC

502

Global Contamination Trends of Persistent Organic Chemicals

22.1.5  Study Objectives Time-series studies of environmental contaminants are usually conducted to understand trends in exposure levels in particular organisms or regions. While time-series studies have indicated increasing trends in PBDE concentrations in both abiotic and biotic matrices, considerable data gaps remain in the current monitoring studies such as time-trend studies of HBCDs. Most HBCD studies have been conducted in Europe and the United States, while very few studies have focused on the Asia-Pacific region. There is hence an urgent need to provide a better understanding of global HBCD concentrations in the environment. This paper describes temporal trends in PBDE and HBCD concentrations in the Indo-Pacific humpback dolphin (Sousa chinensis) and finless porpoise (Neophocaena phocaenoides), the two resident cetacean species in Hong Kong. The dolphin is generally restricted to the northwestern waters of Hong Kong adjacent to the mouth of the Pearl River, whereas the porpoise is found in eastern waters (Figure 22.2). Both cetacean species are top predators, which have been the subjects of previous monitoring studies [75–77].

Pearl River Estuary

Shenzhen

Hong Kong

(a)

Pearl River Estuary

Shenzhen

Hong Kong

(b)

Figure 22.2  (a) Indo-Pacific humpback dolphin sightings ( ) in Hong Kong and its neighboring waters. (b) Finless porpoise sightings ( ) in Hong Kong coastal waters.

© 2012 by Taylor & Francis Group, LLC

Temporal Trends of Polybrominated Diphenyl Ethers

503

22.2  Materials and Methods 22.2.1  Sample Collection and Preparation Blubber samples of stranded Indo-Pacific humpback dolphins (n = 17) and finless porpoises (n = 33) were collected in Hong Kong between 2002 and 2007 and between 2003 and 2008, respectively. All of the samples were stored in plastic bags and transported to the laboratory for storage at −20°C prior to chemical analysis. Analysis of PBDEs and HBCDs followed previously established methods [26,78] with modifications. Blubber samples were freeze-dried and then homogenized. The homogenized samples were extracted in a Soxhlet apparatus for 10 h with 400 mL of a hexane:dichloromethane (DCM) mixture (1:3 v/v). Before the extraction, PCB 30 and decachlorobiphenyl (DCB) were added to the samples as the surrogate standards. The extract was concentrated and a portion of the sample extract was used for lipid determination by the gravimetric method. The extract was added to a gel permeation chromatography column (GPC; Bio-Beads S-X3, Bio-Rad Laboratories) for lipid removal by an equivalent mixture of DCM in hexane (1:1 v/v) at a flow rate of 5 mL/min. 5 ng of each 13C12-labeled PBDE standards (13C12-labeled BDE 3, BDE 15, BDE 28, BDE 47, BDE 99, BDE 153, BDE 154, BDE 183, BDE 197, BDE 207, and BDE 209) and 10 ng of each 13C12-labeled HBCD standard (α-, β- and γ-13C12labeled HBCD) were spiked to the extract before adding it to the GPC column. The concentrated extract was further purified by elution through 4 g activated silica gel (60 Å´ average pore size) and 5 g activated alumina. PBDEs were eluted with 120 mL DCM:hexane (1:20 v/v) and HBCDs eluted with a DCM:hexane mixture (2:1 v/v) from the silica gel column. 13C12-labeled BDE 139 was added to the eluent as the recovery spike and the volume was further reduced to 0.1 mL prior to gas chromatograph (GC) analysis.

22.2.2  Chemical Analysis Instrumental analysis of PBDEs was performed using a GC (Agilent 7890A) coupled with an MSD (Agilent 5975c) for mono- to deca-brominated diphenyl ethers (BDEs) using electron impact (EI) mode. Lower and higher brominated PBDE congeners were analyzed by 30 m DB-5MS and 15 m DB-5HT columns respectively. Total PBDE (ΣPBDE) concentrations were reported as the sum of the 14 individual, PBDE congeners, and each congener was quantified using the isotope dilution method to their corresponding 13C12-labeled congeners. Quantification of HBCDs followed the analytical method previously established by Tomy et al. [56,79] using an Agilent HP1100 liquid chromatography (Agilent, Palo Alto, CA) equipped with an Applied Biosystems API 2000 triple quadrupole tandem mass spectrometer (MS) coupled with a Turbo IonSpray source operated in negative mode. The three isomers of (α-, β- and γ-) HBCDs were separated chromatographically by a Symmetry-C18 column (2.1 mm i.d. × 150 mm, 3.5 μm, Waters Corp.) equipped with a guard column. The MS/MS analysis was performed using electrospray ionization (ESI) with multiple reaction monitoring mode (MRM). Native α-, β-, and γ-HBCD isomers were quantified by the mean value of the response at two MRM transitions (m/z 640.6 > 81 and 640.6 > 79) corrected against the response of 13C12-labeled HBCDs (m/z 652.6 > 81 and 652.6 > 79). Total HBCD (ΣHBCD) concentrations were reported as the sum of the three individual diastereoisomers quantified. Concentrations of both PBDEs and HBCDs were expressed as ng/g lipid weight (lw).

22.2.3  QA/QC Recoveries of 13C12-labeled HBCDs and PBDEs in all samples were within 60%–120%. The efficiencies of Soxhlet extraction and cleanup procedures were checked prior to the chemical analysis and the recovery rates of 13C12-labeled standards ranged between 70% and 120% (n = 5). Procedural

© 2012 by Taylor & Francis Group, LLC

504

Global Contamination Trends of Persistent Organic Chemicals

blanks were analyzed simultaneously with every batch of five samples to check for potential interferences or contamination and instrumental detection limits (IDLs) were estimated as the average signal of the blanks plus three times the standard deviation of the signals of the blanks. The IDL was 0.02 ng/g lw for mono- to di-BDEs, 0.1 ng/g lw for tetra-BDEs, 0.05 ng/g lw for tri- and penta- to hepta-BDEs, 0.02 ng/g lw for octa- to nona-BDEs, 0.5 ng/g lw for deca-BDE and 0.3 ng/g lw for each diastereoisomer of HBCDs.

22.2.4  Statistical Analysis Concentration comparisons were conducted by student’s t-tests if the data passed normality and equal variance tests; nonparametric Mann–Whitney Rank Sum tests were used otherwise (SigmaStat 3.5). ΣPBDE and ΣHBCD concentrations in Indo-Pacific humpback dolphins and finless porpoises sampled from 1997 to 2001 reported by Isobe et al. [78] were incorporated into the present data set of blubber samples collected from 2002 to 2008 for time-series analyses. Temporal trend analyses were carried out using a simple log-linear regression model as described by Nicolson et al. [80]. Briefly, log-linear regression was performed using annual mean concentrations of ΣHBCDs or ΣPBDEs for dolphins from 1997 to 2007 and porpoises from 2000 to 2008. A 3 year moving average smoothing function was fitted to the annual median concentrations to investigate possible nonlinear trend components and was tested by means of ANOVA. To minimize the influence of differences between the sexes due to maternal transfer and possible age-related differences, adult males with body lengths greater than 200 and 120 cm for dolphins and porpoises, respectively, were selected for temporal trend analyses. Statistical analysis was conducted using Prism 2.01 and SigmaStat 3.5. Statistical significance was accepted at p < 0.05.

22.3  Results and Discussion 22.3.1  Concentrations of PBDEs and HBCDs in Cetaceans HBCDs were detected in all 50 cetacean blubber samples at concentrations ranging from 32 to 519 and from 4.1 to 501 ng/g lw for Indo-Pacific humpback dolphins and finless porpoises, respectively. The dominant congener was α-HBCD in all samples analyzed. The detection frequencies of β- and γ-HBCD in the samples analyzed were 54% and 60%, respectively. The average ΣPBDE concentrations ranged from 1113 to 3590 ng/g lw in the two species, and were approximately one order of magnitude greater than the levels of ΣHBCDs in blubber samples of both cetacean species. Concentrations of HBCDs and PBDEs in dolphins were significantly greater than those in porpoises (p < 0.05), suggesting higher exposure levels of the two contaminant groups in the northwestern waters of Hong Kong than in the eastern waters.

22.3.2  Temporal Trend of PBDEs and HBCDs in Cetaceans A positive linear temporal trend in ΣHBCD concentrations was found in dolphin samples when loglinear regression analysis of yearly median concentrations was used (Pearson r = 0.67, p = 0.047). The linear regression equation obtained for dolphins (y = 0.090x + 1.6) indicated that ΣHBCD concentrations would double from 1997 to 2017, assuming a constant rate of HBCD usage and release [81]. The world market demand for HBCDs has increased during the past decade, while demand for traditional BFRs such as PBDEs is decreasing because of controls on their production and use. The usage of HBCDs as alternatives for PBDE formulations is projected to increase in any region with rapid economic development and industrial activities, such as the PRD; there is currently no restriction or control on HBCD production and use in China. Therefore, it is not surprising that elevated trends of HBCDs were observed in Indo-Pacific humpback dolphins in the present study, because this species inhabits the northwestern waters of Hong Kong located downstream of the PRD. In contrast, no

© 2012 by Taylor & Francis Group, LLC

Temporal Trends of Polybrominated Diphenyl Ethers

505

significant time trend in HBCD concentrations was detected from 2000 to 2008 for finless porpoises whose habitats are generally located in the less polluted eastern waters of Hong Kong. Concentrations of ΣPBDEs in both dolphins and porpoises, in contrast, showed no significant trend for the corresponding sampling years when tested by log-linear regression and 3 year moving averages. However, between-year variation of ΣPBDEs exhibited similar patterns between the two cetacean species with peak concentrations occurring in 2003–2004 followed by relatively steady levels after 2006. Similar patterns were found in guillemot (Uria aalge) eggs from the Baltic Sea showing a positive trend from 1970 to the middle of the 1980s, followed by a decrease up to 2001 [67]. These observations could be partly explained by effective regulation and control of PBDE usage in countries around the world.

22.3.3  Global Trends of PBDE and HBCD Concentrations in Marine Mammals The current knowledge of the temporal trends of PBDE and HBCD levels in marine mammals is limited to North America (United States), Europe (United Kingdom), Asia-Pacific (Japan), and the Arctic (Canada, Greenland, and Norway), in addition to the current study location, Hong Kong. In general, increasing trends in HBCDs concentrations have been found in marine mammals from these regions, while no continuous decreasing trends of PBDEs have been observed. These results agree with the findings of the present study. 22.3.3.1  Japan Studies on temporal trends of PBDEs in marine mammals are scarce in the Asia-Pacific region; the majority of the available information comes from Japan. Generally, the last few decades witnessed a significant increase in PBDE levels in marine mammals; concentrations in the 1970s and 1980s were 1–2 orders of magnitude higher than those in recent years [61]. In northern fur seals (Callorhinus ursinus), an approximately 150-fold elevation of PBDE concentrations was observed between 1972 and 1994, and then the levels dropped by approximately half between 1994 and 1998 [82]. A similar increase in average PBDE concentrations from 13 ng/g lw in 1978 to 640 ng/g lw in 2003 was observed in striped dolphins (Stenella coeruleoalba) [83]. PBDE levels in melon-headed whales (Peponocephala electra) also showed a 10-fold increase from 1982 to 2006 [84]. In fur seals and striped dolphins, PBDE concentrations peaked in the early 1990s and appeared to decline afterward. Among PBDE congeners, BDE 47 was the most abundant congener in all the marine mammal species across all sampling years. There was, however, temporal variation among congener profiles, suggesting the use of different PBDE commercial formulations during different time periods. In northern fur seals, the ratios of BDE 153, BDE 154, and BDE 183 to ΣPBDEs increased after 1972, while those of some lower brominated congeners decreased. In melon-headed whales, hexabrominated congeners comprised a higher proportion of ΣPBDEs in 2006 than in 1982. These observations might arise from the ban of the Penta-BDE mixture in 1990 and subsequent shift toward the use of higher brominated formulations, namely Octa-BDE (withdrawn in 2000) and Deca-BDE in Japan [14]. Like PBDEs, HBCD concentrations demonstrated a sharp rise in marine mammals from Japan during recent decades. In northern fur seals, HBCD concentrations increased from

Suggest Documents