Polybrominated Diphenyl Ethers (PBDEs) in Biota Representing ...

2 downloads 24 Views 3MB Size Report
May 6, 2008 - dominance in many biotic samples (17). ANOVA tests on the experimental data were conducted using the Analysis ToolPak of Microsoft Excel.

Environ. Sci. Technol. 2008, 42, 4331–4337

Polybrominated Diphenyl Ethers (PBDEs) in Biota Representing Different Trophic Levels of the Hudson River, New York: From 1999 to 2005 K A N G X I A , * ,†,⊥ M I N G B O L U O , ⊥ C H R I S T I N A L U S K , † K E V I N A R M B R U S T , †,⊥ LAWRENCE SKINNER,‡ AND RONALD SLOAN‡ Mississippi State Chemical Laboratory, Department of Chemistry, P.O. Box CR, Mississippi State, Mississippi 39762, and New York State Department of Environmental Conservation, 625 Broadway, Albany, New York 12233

Received December 19, 2007. Revised manuscript received March 11, 2008. Accepted March 18, 2008.

It has been hypothesized that a principal route of human exposure to polybrominated diphenyl ethers (PBDEs), used as flame retardants, is through fish consumption. Between 1999 and 2005 PBDE-47, -99, -100, -153, and -154 were analyzed in 3797 biological samples of 33 species of the Hudson River, New York. Approximately 98.4% of the samples contained PBDEs between 0.5 and 37 169 ng g-1 lipid, with a median concentration of 772 ng g-1 lipid. Yearly median ΣPBDE concentrations fluctuated. Samples from river miles 112 and 153 contained higher ΣPBDEs than those from other locations of the river. The 7-year median ΣPBDE concentrations were the highest in large carnivorous fishes and the lowest in insects. The median abundance of congener PBDE-47 decreased from 80% to 63% with decreasing levels of ΣPBDEs in the samples, while an increase from 2% to 23% was observed for PBDE-99. The median abundance of other congeners did not change with concentrations of ΣPBDEs. Positive-, negative-, and no-correlation between ΣPBDE concentrations and fish weight were observed for different species and for the same species from different locations of the river. The sources of PBDE contamination, diet, metabolic activity, and sediment chemistry might affect the levels of PBDEs in a fish.

Introduction Polybrominated diphenyl ethers (PBDEs) have been widely used in the United States and worldwide as flame retardants (1). Due to increasing concern of endocrine-disrupting effects (2, 3) and developmental neurotoxicity (4, 5), the use of pentaPBDE and octa-PBDE, which are more toxic than PBDEs with higher degree of bromination (6, 7), has been recently prohibited in eight U.S. states. Their manufacture in the United States was voluntarily phased out at the beginning of 2005. However, a large number of products containing * Corresponding author phone: 662-325-5896; fax: 662-325-7807; e-mail: [email protected] † Mississippi State Chemical Laboratory. ⊥ Department of Chemistry. ‡ New York State Department of Environmental Conservation. 10.1021/es703049g CCC: $40.75

Published on Web 05/06/2008

 2008 American Chemical Society

penta-PBDEs and octa-PBDEs are still currently in use, resulting in the continued release of those compounds into the environment from the use, recycling, and disposal of those products (8, 9). In addition, deca-PBDE continues to be permitted for use worldwide and its production is on the rise (10). There is limited information on how deca-PBDE transformation affects the input of lower brominated PBDEs into the environment (11, 12). Biological degradation of decaPBDE to the lesser-brominated PBDE congeners under anaerobic conditions was recently reported (12). PBDEs are considered candidate persistent organic pollutants (POPs) by the Stockholm Treaty (13). Penta-PBDEs and octa-PBDEs bioaccumulate and biomagnify in biological samples due to their high lipophilicity and persistence in environmental media (14). For pike, perch, and roach from the Baltic Sea the biomagnification potential was positively correlated with degree of bromination for tri- to penta-PBDEs and reached a maximum for penta-PBDEs, while the opposite trend was found for hexa- to hepta-PBDEs (14). Biomagnification did not occur for octa-, nona-, and deca-PBDEs in this study. Large molecular size and high molecular weight of PBDEs with higher degree of bromination might restrict their accumulation due to inefficient dietary uptake (14). However, rapid biotransformation via debromination of decaPBDEs was observed in studies with rainbow trout (15) and salmon (16), and may contribute to the relatively lower levels of PBDE-209 in many biological samples compared to those for PBDEs with a lesser degree of bromination (17). Between 1970 and 2000, the concentrations of ΣPBDEs in marine mammals, birds, and humans have doubled approximately every 4 to 7 years, while the temporal change for ΣPBDEs in fishes was not as systematic during the same time period (17, 18). To date, the most frequently reported PBDE congeners in environmental samples are PBDE-47, PBDE-99, PBDE-100, PBDE-153, and PBDE-154. Congener PBDE-47 is most dominant in a majority of biological samples. Due to analytical difficulties, congener PBDE-209 has not been included as an analyte of interest in many published reports. In the few investigations that measured PBDE-209, its levels were less significant in biological samples but more prevalent in sediment and sewage samples than other lower brominated congeners (17, 18). Most of the aquatic biota investigations published to date have been conducted on samples collected from the North Atlantic Ocean and Europe. Although there are several reports on levels of PBDEs in freshwater biota from North American lakes (19, 20) and rivers (21–23), most of the measurements were made on fish occupying the highest trophic level in the food web of those ecosystems. No previous studies have examined both the temporal and spatial patterns of PBDEs in biota at different trophic levels within a river ecosystem in North America. This study examines these relationships within the Hudson River for select PBDE congeners.

Materials and Methods A total of 3797 biological samples representing 33 species of different trophic levels (see Tables S1 and S2 in Supporting Information) were collected by the staff of the New York State Department of Environmental Conservation during the same season each year from 1999 to 2005 at 16 locations of the Hudson River, New York (see Figure S1 in Supporting Information). The locations of the sampling sites stretched from river mile (RM) 13 to 204 (the distance in miles along the Hudson River north of the southern tip of Manhattan Island in New York City). Samples consisted of single fish for VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY



FIGURE 1. Box plot (log) concentrations of ΣPBDEs in biota of the Hudson River from 1999 to 2005. The numbers of data points are given in parentheses. Dashed lines indicate average values. The line within each box, the lower and upper boundaries, the error bars below and above the box, and the circles below and above the error bars mark the median (50th percentile), the 25th and 75th, 10th and 90th, 5th and 95th percentiles, respectively. large species and composites of multiple individuals for small species. Weight, size, and sex of each sample were documented. Samples were packed on ice upon collection, frozen at -20 °C after returning to the laboratory, and shipped frozen to the Mississippi State Chemical Laboratory for analysis. Procedures for sample processing, analytes extraction, cleanup, analysis, and confirmation can be found in the Supporting Information (Section S2). Congeners PBDE-47, -99, -100, -153, and -154 were analyzed because of their dominance in many biotic samples (17). ANOVA tests on the experimental data were conducted using the Analysis ToolPak of Microsoft Excel. Results were considered significantly different at p < 0.01 unless noted otherwise (24).

Results and Discussion Occurrence of PBDEs in Biota of the Hudson River. About 98.4% of the 3797 biota samples (see Table S1 in Supporting Information) collected from 1999 to 2005 contained at least one of the five PBDE congeners at a concentration above the quantitation limit (0.5 ng g-1 lipid), with the highest concentration reaching 37 169 ng g-1 lipid in a striped bass sample. This concentration range is much wider than that (6.3-7200 ng g-1 lipid) detected for the European and North America fishes sampled between 1987 and 2000 (17, 18). The highest total PBDE concentration in edible fish tissue recently reported was 47 900 ng g-1 lipid in a carp collected from the Hyco River, Virginia (21). Of all the samples collected during the 7-year period, the ΣPBDE levels were between 10 000 and 37 169 ng g-1 lipid for 81 samples (2.1%), between 5000 and 10 00 ng g-1 lipid for 171 samples (4.5%), between 1000 and 5000 ng g-1 lipid for 1303 samples (34.3%), between 500 and 1000 ng g-1 lipid for 980 samples (25.8%), and between 0.5 and 500 ng g-1 lipid for 1202 samples (31.7%). Sixty samples (1.6%) contained no detectable PBDEs. The median ΣPBDE concentration for all the samples was 772 ng g-1 lipid. Figure 1 shows that the median levels of ΣPBDEs in the biological samples collected each year fluctuated with time between 1999 and 2005. The highest median concentration of ΣPBDEs appeared in year 2000, reaching 1915 ng g-1 lipid, while the lowest was about 500 ng g-1 lipid, both in 2002 and 2005. Significant accumulation of PBDEs at the top trophic level of the food web in the Hudson River is evident. Figure 2 shows the 7-year median concentration of ΣPBDEs in large carnivores (average adult size >30 cm) is close to 3 times 4332



FIGURE 2. Box plot (log) concentrations of ΣPBDEs in organisms of different trophic levels of the Hudson River. The number in each box indicates the median concentration of ΣPBDEs. Dashed lines indicate average values. The line within each box, the lower and upper boundaries, the error bars below and above the box, and the circles below and above the error bars mark the median (50th percentile), the 25th and 75th, 10th and 90th, 5th and 95th percentiles, respectively. The numbers in centimeters on the x-axis indicate the average lengths of adult fishes in each trophic level category (see Table S2 and Section S2 in Supporting Information for trophic level classification discussion). those in smaller fishes and in invertebrates and more than 8 times those in insects. Large carnivores in general occupy the highest level of the food web in an aquatic ecosystem (see Table S2 and Section S2 in Supporting Information) and, therefore, may accumulate PBDEs when consuming other organisms contaminated with PBDEs. Although small carnivores, invertivores, and large and small invertivores/ detritivores often feed on invertebrates, their 7-year median concentrations of ΣPBDEs are similar to that in invertebrates (Figure 2). The lowest 7-year median concentrations of ΣPBDEs are tested in insects, which generally occupy the lowest trophic level of an aquatic ecosystem and are consumed by many fishes and invertebrates. Carnivorous insects such as hellgrammites, dragonfly, and damselfly most often feed on other insects, they also occasionally feed on invertebrates and even juvenile fish. The 7-year median concentration of ΣPBDEs in carnivorous insects is slightly higher than that in herbivorous insects such as caddisfly. Congener PBDE-47 was the most abundant in 92% of the samples containing detectable levels of ΣPBDEs. Figure 3 shows the abundance of PBDE-47 decreased with decreasing concentrations of ΣPBDEs in the biological samples. For half of the samples that contained ΣPBDEs higher than 10 000 ng g-1 lipid, the contribution of PBDE-47 to the total is more than 80%. However, the median contribution of PBDE-47 relative to total decreased to 63% for samples containing ΣPBDEs less than 500 ng g-1 lipid. The opposite trend was observed for the contribution of PBDE-99. The median abundance value for PBDE-99 was only 2% in samples containing ΣPBDEs at the highest concentration range, while it was 23% in samples containing ΣPBDEs at less than 1000 ng g-1 lipid (Figure 3). The contribution of congeners PBDE100, -153, and -154 to the total did not vary much with levels of ΣPBDEs detected in the samples. Hale et al. (21) observed a lack of accumulation of PBDE99 (0.1% of the total) in a wild-caught carp that had ΣPBDEs at 47 900 ng g-1 lipid, while as much as 30.3% of the total PBDEs were PBDE-99 in a channel catfish containing ΣPBDEs at 2130 ng g-1 lipid. The data presented in Figure 3 suggest

FIGURE 3. Median percentage contribution of each PBDE congener to the total in organisms containing ΣPBDEs at different concentration ranges. that PBDE-47 and PBDE-99 exhibit differential uptake, metabolism, and excretion for organisms of different trophic levels. Investigations of dietary uptake of PBDEs by carp (25) indicated that production of PBDE-47 in carp tissues occurred as a result of debromination of higher-brominated compounds, possibly PBDE-99. Based on their dietary exposure study on salmon, Isosaari et al. (26) implied that the reason for the lower accumulation of PBDE-99 in fish, relative to PBDE-47, was not only due to debromination but also to preferential excretion of PBDE-99. Temporal and Spatial Trends of PBDEs in Various Aquatic Species. A total of 328 catfish samples including channel catfish, white catfish, yellow bullhead, and brown bullhead were collected from 10 locations of the Hudson River between 1999 and 2005. Only two samples (brown bullhead) had PBDEs below the detection limit. The concentrations of ΣPBDEs in the catfish samples ranged from 20 to 13 229 ng g-1 lipid. The highest concentration of ΣPBDEs was detected in a white catfish caught at RM27. The median ΣPBDE concentration in the catfish tested was 1097 ng g-1 lipid. White catfish had the highest median ΣPBDE concentration at 2225 ng g-1 lipid, followed by channel catfish, brown bullhead, and yellow bullhead, with median ΣPBDE concentrations of 1521, 880, and 787 ng g-1 lipid, respectively. Figure 4 shows that the catfish caught at RM27, 60, 112, and 153 had much higher median ΣPBDE concentrations, ranging from 2003 to 2816 ng g-1 lipid, than those caught at other locations (105-1870 ng g-1 lipid). Temporal changes in ΣPBDE concentrations in the catfish samples were not obvious at any location sampled. PBDEs were detected in all of the 1039 striped bass and white perch collected from 10 locations of the Hudson River during the 7-year period. Both striped bass and white perch are extremely popular with many anglers and belong to the temperate bass family (27). The concentrations of ΣPBDEs in the bass samples ranged from 104 to 37 169 ng g-1 lipid, with a median concentration of 1568 ng g-1 lipid. The highest ΣPBDE concentration was detected in a striped bass collected at RM153. In general, white perch contained higher PBDE concentrations than striped bass. The median ΣPBDE concentrations were 3685 and 1322 ng g-1 lipid for white perch and striped bass, respectively. Figure 4 shows that striped bass and white perch from locations further north on the Hudson River had higher levels of PBDEs than those caught at locations closer to the New York City. High levels

of PBDEs were detected each year in both species collected at RM153 (Figure 4). The median ΣPBDE concentrations in the two species from this location varied from 6066 ng g-1 lipid in samples from year 2002 to 1611 ng g-1 lipid in samples from year 2005. At RM112, high PBDE levels were detected in samples collected in 2000 and 2005, reaching median concentrations of 4532 and 5232 ng g-1 lipid, respectively. A systematic temporal change in PBDE levels for both temperate bass species from the Hudson River was not observed (Figure 4). Only 4 yellow perch samples out of the 349 fishes (yellow perch and walleye) from the large perch family contained PBDEs below the detection limit. Compared to catfish and temperate bass, fishes from the large perch family had slightly lower median ΣPBDE concentration, at a level of 662 ng g-1 lipid for the 349 fishes investigated. The ΣPBDE concentrations in those fishes ranged from 57 to 27 103 ng g-1 lipid. The highest ΣPBDEs concentration was detected in a yellow perch caught at RM153. The median ΣPBDE concentration in walleye from the Hudson River during the 7-year period was 2613 ng g-1 lipid, four times that in yellow perch. Similar to observations for fishes from the catfish family and the temperate bass family, yellow perch and walleye from RM153 contained high levels of PBDEs with median concentrations ranging from 2143 to 9108 ng g-1 lipid (Figure 4). In addition, yellow perch and walleye caught at locations RM112, 76, and 60 also contained high levels of PBDEs. The median ΣPBDE concentrations in both bass species collected each year fluctuate with time between 1999 and 2005 (Figure 4). A total of 1224 fish belonging to the sunfish family (including redbreast sunfish, pumpkinseed, bluegill, rock bass, smallmouth bass, and largemouth bass) were collected from 11 locations of the Hudson River (Figure 4). PBDEs were detected in 1214 samples, ranging from 11 to 32 219 ng g-1 lipid, with a median concentration of 664 ng g-1 lipid. Both largemouth bass and smallmouth bass contained much higher PBDEs than the other sunfish species. The median ΣPBDE concentrations in the largemouth bass and smallmouth bass were 1913 and 1785 ng g-1 lipid, respectively, while those in the bluegill, pumpkinseed, redbreast sunfish, and rock bass were 226, 438, 570, and 595 ng g-1 lipid, respectively. Compared to other sunfish species, both largemouth bass and smallmouth bass are at a higher level of the food chain, resulting in greater bioaccumulation of PBDEs. Although decreasing sharply with time from 2000 to 2003, VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY



FIGURE 4. Median ΣPBDE concentrations in organisms of the catfish family, temperate bass family, large perch family, sunfish family, and minnow family collected at different locations of the Hudson River between 1999 and 2005. River mile is the distance (miles) on the river north of the southern tip of Manhattan Island in New York City. the levels of PBDEs in sunfish collected at RM153 tend to be higher than those in the sunfish collected at most of the other locations (Figure 4). Temporal changes in PBDE concentrations in the sunfishes were not obvious except as noted at RM153. Two hundred six fish samples representing the minnow family (including golden shiners, spottail shiners, bluntnose minnows, fathead minnows, and fallfish) were collected from six locations on the Hudson River from 2003 to 2005 (Figure 4). Only 4 of the 206 fish samples had nondetectable PBDE concentrations. The ΣPBDE concentrations ranged from 10 to 5348 ng g-1 lipid, with a median concentration of 388 ng g-1 lipid. Of the five minnow species investigated, fathead minnows had the lowest median ΣPBDE concentration (32 4334



ng g-1 lipid) while spottail shiners had the highest median level (508 ng g-1 lipid). Once again, fishes collected at RM153 had the highest PBDEs compared to those from other locations. The median ΣPBDE concentrations in the fishes collected at RM76 and 112 decreased to about half of their level from 2003 to 2005, while PBDE levels remained constant during the 3-year period in minnows from other locations (Figure 4). The temporal and spatial changes of PBDEs in carp, American eel, white suckers, blue crabs, and blueback herring are discussed in the Supporting Information (S4). No obvious temporal trends were observed for ΣPBDE concentrations in those species.

FIGURE 5. Median percentage composition of each PBDE congener in individual fish species. A total of 439 biological samples including amphibians, invertebrates, and insects (see Table S1 in Supporting Information) were collected at RM176 and 190 between 2002 and 2005. Since there were no samples taken of macroinvertebrates south of RM176, there is no real capability to examine spatial relationships for PBDEs in this group of species from along the length of the river. About 8.9% of the samples contained nondetectable PBDEs. The highest concentration was detected in a snail sample at 11 905 ng g-1 lipid. The caddisflies had the lowest median ΣPBDE concentration at 112 ng g-1 lipid, while snails had the highest median concentration at 523 ng g-1 lipid. In general, temporal changes in ΣPBDE concentrations were not observed during the 7-year investigation. There is a spatial difference in PBDE concentrations across most species or species classes in the two top trophic levels. There are consistently lower PBDE levels in fish taken north of RM153 than in samples taken from downstream locations. RM153 is an apparent hot spot for PBDEs, likely due to increased exposures which are consequent to the increased human population in the area and the associated human uses of products containing PBDEs. RM153 is the location of the Federal Dam which reaches from Troy on the east shore to Green Island on the west shore and is generally considered the approximate northern extent of the Capital District (Albany, Troy, Schenectady, Rensselaer area). The Capital District is the first major populated area on the river thus would be expected to have the first major PBDE sources to the river. Also, RM153 is the beginning of the tidal portion of the river and it is generally the upstream limit for most migratory fish. Distribution of PBDE Congeners in Various Aquatic Species. The distribution of the five PBDE congeners in the biological samples tested did not exhibit systematic temporal and spatial changes. The PBDE congener distribution did vary among the aquatic species investigated (Figure 5). The median values of the PBDE-47 contribution to the total in fishes belonging to the catfish family were between 53 and 69%, and the contribution of PBDE-99 to the total was slightly less and decreased with increasing PBDE-47 contribution. Hale et al. (21) reported 47% of the total PBDEs in Virginia

channel catfish were attributed to PBDE-47, a value smaller than reported in the current investigation. The contribution of PBDE-100 to the total in the catfish species remained constant at a median level of around 14%. The median contributions of PBDE-153 and PBDE-154 to the total in the catfish species were below 2%. Compared to brown bullhead and channel catfish, the contribution of PBDE-47 to total was significantly higher (P < 0.01) in both white catfish and yellow bullhead (Figure 5), perhaps due to their higher capability of accumulating PBDE-47 than the other two catfish species or different exposure to food sources. The distributions of PBDE congeners in walleye and yellow perch were similar to those in the catfish species (Figure 5). Striped bass and white perch are from the same family and exhibit similar PBDE congener distribution (Figure 5), with median PBDE-47 contribution to the total around 80%. It is apparent that in addition to accumulating PBDE-47, both bass species tend to preferentially accumulate PBDE100 over PBDE-99, PBDE-153, and PBDE-154. A similar observation was made by Hale et al. (21) for Virginia striped bass. American eels contained PBDE-47 at a median value of 69% of the total, followed by PBDE-100 with a median composition of 22% (Figure 5). The median contribution of PBDE-99 to total in the American eels tested was insignificant, similar to that of PBDE-153 and -154. Lepom et al. (28) observed a similar PBDE distribution pattern for eels caught from the Elbe River in Germany. PBDE-47 and PBDE-100 were the only two congeners detected in carp caught in the Hudson River (Figure 5). The median contribution of PBDE47 to total was 90%. A lack of accumulation of PBDE-99 and PBDE-153 and -154 was also observed for wild-caught Virginia carp (21) and in those exposed in laboratory tests (25). Although largemouth bass and smallmouth bass contained much higher PBDEs than other fish species from the sunfish family, the PBDE congener distributions in all the sunfish species were remarkably similar (Figure 5). The median contributions of PBDE-47 to total were in the range of 53-61%, followed by PBDE-99 at 20-36%, PBDE-100 at 8-14%, and PBDE-153 and -154 at 0-7%. The PBDE congener distribution pattern in sunfish species observed in this study was similar to those made earlier in sunfishes caught from VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY



a Missouri lake (20) and a Virginia river (21). A large portion of the PBDEs detected in fishes belonging to the minnow family was PBDE-47, with a median contribution ranging from 63% in the fathead minnow to 93% in the fallfish (Figure 5). The contribution of PBDE-99 to the total was low. Many fishes had PBDE-99 below the detection limit. The contributions of PBDE-100 to the total in spottail shiner and fathead minnow were 15% and 22%, respectively, more significant (P < 0.01) than those in other members of the minnow family. Compared to other minnows, the fathead minnow accumulated significant levels of PBDE-153 and -154, contributing 14% of the total. Both PBDE-47 and PBDE-99 contributed significantly to the total in tessellated darter and blue crab (Figure 5), but only small portions of the total PBDEs in both species were congeners 100, 153, and 154. About 90% of the total PBDEs in the white sucker were congener PBDE47, while PBDE-100 contributed only about 8% to the total. Congeners PBDE-99, -153, and -154 were rarely detected in the white sucker. The median contributions of congeners PBDE-47, -99, and -100 to the total in blueback herring were 70%, 22%, and 8%, respectively (Figure 5). PBDE-153 and -154 were below the detection limit in all blueback herring tested. Similar congener distribution patterns were also observed for herrings caught along the Swedish coast of the Baltic Sea (29). Congeners PBDE-153 and -154 were not detected while PBDE-47, -99, and -100 were dominant in amphibians, invertebrates, and insects of the Hudson River (Figure 5). This might contribute to low levels of detections of PBDE153 and -154 in organisms at higher trophic levels. The median congener contributions to total PBDEs followed the order of PBDE-47 > PBDE-99 . PBDE-100 for hellgrammites, damselfly, and caddisfly, while in crayfish the order was PBDE-47 > PBDE-99 > PBDE-100. Congener PBDE-47 had the highest abundance in snails followed by PBDE-100 and -99 at slightly lower levels. The congener distribution patterns were almost identical for dragonflies and clams with median contributions of PBDE-100, -47, and -99 to total of approximately 43%, 34%, and 23%, respectively. The median contributions of PBDE-100 and -47 were 89% and 11%, respectively, in bullfrog tadpoles, with PBDE-99 levels below the detection limit. However, significant levels of PBDE-99 appeared in adult bullfrogs, contributing to 43% of the total PBDEs detected. The median contribution of PBDE-100 was lower while the PBDE-47 contribution was greater in adult bullfrogs compared to those in the bullfrog tadpoles. The significantly different congener distribution pattern between the tadpoles and adult bullfrogs is possibly due to their dramatically different feeding habits, and consequently PBDE exposure. Tadpoles feed on planktonic species of both animal and plant origin from within the water column where they live. During metamorphosis, there is a gradual dietary change to large food organisms. In contrast, adult bullfrogs temporarily range from their wetland habitats on a daily basis and feed on earthworms, insects, and a wide array of other organisms, including mice and shrews. Correlation of ΣPBDEs with Fish Characteristics. Compared to studies published by others, this study has examined PBDEs in the largest population of aquatic organisms to date. For most fish species, except striped bass, no significant concentration differences were found for ΣPBDEs in male and female fishes with similar weight. Smaller male striped bass contained higher levels of ΣPBDEs than female fish of similar weight, while the difference was not as significant for larger fish (see Figure S2 in Supporting Information). Similarly no significant concentration differences between male and female fishes were observed for PBDEs in 11 lake trout (30) or for PCBs and chlorinated pesticides in 130 perch samples (31). 4336



Most striped bass are migratory, although a very small portion of the population of striped bass within the Hudson River during spring migrations are believed to be resident to the river throughout their life span. For males, once they have returned to the ocean after spring spawning migrations, there is likely growth dilution of the accumulated PBDE concentrations thereby, over time, displaying smaller concentrations in larger individuals. For females, there may have been transfer of PBDEs from the body of the female to eggs prior to or during their entry into the river. During preparation for and the period of spawning, there is highly active development of the reproductive products with massive transfer of fats/oils (and a portion of their contaminants) from the body to the eggs. This may cause the concentrations of chemical residues in the body to be reduced in relation to PBDE concentrations in males of the same size. Other studies documented positive correlations between concentrations of hydrophobic contaminants such as PBDEs and fish weight (21, 31). Our investigation suggests no consistent relationships (positive, negative, or otherwise) between concentration of ΣPBDEs and weight of fish of the Hudson River. Some examples for brown bullhead, striped bass, yellow perch, largemouth bass, and smallmouth bass are shown in Figure S2 of Supporting Information. The correlations between the levels of ΣPBDEs in fishes and their weight varied among species and locations of the Hudson River (see S4 in the Supporting Information for detailed discussion). The wide variations in food preferences and the metabolic differences among fish species, or within a species at different ages, might contribute to the observed variation between the levels of ΣPBDEs in aquatic organisms and their physical characteristics. In addition, differences in PBDE sources, water, and sediment chemistry at different locations along the river might also impact PBDEs uptake and accumulation in aquatic organisms.

Acknowledgments We thank the staff of the New York State Department of Environmental Conservation for sample collection and the staff of the Mississippi State Chemical Laboratory for sample analysis. We also thank Dr. Steve Miranda of the Department of Wildlife & Fisheries at Mississippi State University for providing information on trophic level classification for aquatic organisms.

Supporting Information Available Detailed information on sampling locations, species of organisms collected, and sample processing, extraction, cleanup, and analysis. This material is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) van Esch, G. J. Environmental Health Criteria 162: Brominated diphenyl ethers; IPCS International Program on Chemical Safety, United Nations Environment Program, International Labor Organization, World Health Organization, 1994; available at http://www.inchem.org/documents/ehc/ehc/ehc162.htm. (2) Meerts, I. A. T. M.; van Zanden, J. J.; Luijks, E. A. C.; van LeeuwenBol, I.; Marsh, G.; Jakobsson, E.; Bergman, Å.; Brouwer, A. Potent competitive interactions of some brominated flame retardants and related compounds with human transthyretin in vitro. Toxicol. Sci. 2000, 56, 95–104. (3) Kester, M. H.; Bulduk, S.; van Toor, H.; Tibboel, D.; Meinl, W.; Glatt, H. Potent inhibition of estrogen sulfotransferase by hydroxylated metabolites of polyhalogenated aromatic hydrocarbons reveals alternative mechanism for estrogenic activity of endocrine disrupters. J. Clin. Endocrinol. Metab. 2002, 87, 1142–1150. (4) Eriksson, P.; Jakobsson, E.; Fredriksson, A. Brominated flame retardants: a novel class of developmental neurotoxicants in our environment. Environ Health Perspect. 2001, 109, 903–908. (5) Viberg, H.; Fredriksson, A.; Eriksson, P. Neonatal exposure to the brominated flame retardant 2,2′,4,4′,5-pentabromodiphenyl





(10) (11)







ether causes altered susceptibility in the cholinergic transmitter system in the adult mouse. Toxicol. Sci. 2002, 67, 104–107. Hooper, K.; McDonald, T. A. The PBDEs: an emerging environmental challenge and another reason for breast-milk monitoring programs. Environ. Health Perspect. 2000, 108, 387– 392. Siddiqi, M. A.; Laessig, R. H.; Reed, K. D. Polybrominated diphenyl ethers (PBDEs): new pollutants-old diseases. Clin. Med. Res. 2003, 1, 281–290. Alcock, R. E.; Sweetman, A. J.; Prevedouros, K.; Jones, K. C. Understanding levels and trends of BDE-47 in the UK and North America: an assessment of principal reservoirs and source inputs. Environ. Int. 2003, 29, 691–698. Wu, N.; Herrmann, T.; Paepke, O.; Tickner, J.; Hale, R.; Harvey, E.; La Guardia, M.; McClean, M. D.; Webster, T. F. Human exposure to PBDEs: associations of PBDE body burdens with food consumption and house dust concentrations. Environ. Sci. Technol. 2007, 41, 1584–1589. BSEF. The Bromine Science and Environmental Forum, 2007; available athttp://www.bsef.com/. Hansen, B. G., Munn, S. J., de Bruijn, J., Luotamo, M., Pakalin, S., Berthault, F., Vegro, S., Pellegrini, G., Allanou, R., Scheer, S., Eds. EUR 20402 EN-European Union Risk Assessment Report bis(pentabromophenyl) Ether, Volume 17; European Commission: Luxembourg, 2002. Gerecke, A. C.; Hartmann, P. C.; Heeb, N. V.; Kohler, H. E.; Giger, W.; Schmid, P.; Zennegg, M.; Kohler, M. Anaerobic degradation of decabromodiphenyl ether. Environ. Sci. Technol. 2005, 39, 1078–1083. Pozo, K.; Harner, T.; Wania, F.; Muir, D. C. G.; Jones, K. C.; Barrie, L. A. Toward a global network for persistent oganic pollutants in air: results from the GAPS Study. Environ. Sci. Technol. 2006, 40, 4867–4873. Burreau, S.; Zebu ¨ hr, Y.; Broman, D.; Ishaq, R. Biomagnification of polychlorinated biphenyls (PCBs) and polybrominated diphenyl ethers (PBDEs) studied in pike (Esox lucius), perch (Perca fluviatilis) and roach (Rutilus rutilus) from the Baltic Sea. Chemosphere 2004, 55, 1043–1052. Kierkegaard, A.; Balk, L.; Tja¨rnlund, U.; de Wit, C. A.; Jansson, B. Dietary uptake and biological effects of decabromodiphenyl ether in rainbow trout (Oncorhynchus mykiss). Environ. Sci. Technol. 1999, 33, 1612–1617. Boon, J. P.; van Zanden, J. J.; Lewis, W. E.; Zegers, B. N.; Goksøyr, A.; Arukwe, A. The expression of CYP1A, vitellogenin and zona radiata proteins in Atlantic salmon (Salmo salar) after oral dosing with two commercial PBDE flame retardant mixtures: Absence of short-term responses. Mar. Environ. Res. 2002, 54, 725–728. Hites, R. A. Polybrominated diphenyl ethers in the environment and in people: a meta-analysis of concentrations. Environ. Sci. Technol. 2004, 38, 945–956.

(18) de Wit, C. A. An overview of brominated flame retardants in the environment. Chemosphere 2002, 46, 583–624. (19) Manchester-Neesvig, J. B.; Valters, K.; Sonzogni, W. C. Comparison of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in Lake Michigan Salmonids. Environ. Sci. Technol. 2001, 35, 1072–1077. (20) Dodder, N. G.; Strandberg, B.; Hites, R. A. Concentrations and spatial variations of polybrominated diphenyl ethers and several organochlorine compounds in fishes from the Northeastern United States. Environ. Sci. Technol. 2002, 36, 146–151. (21) Hale, R. C.; La Guardia, M. J.; Harvey, E. P.; Mainor, T. M.; Duff, W. H.; Gaylor, M. O. Polybrominated diphenyl ether flame retardants in Virginia freshwater fishes (USA). Environ. Sci. Technol. 2001, 35, 4585–4591. (22) Rice, C. P.; Chernyak, S. M.; Begnoche, L.; Quintal, R.; Hickey, J. Comparisons of PBDE composition and concentration in fish collected from the Detroit River, MI and Des Plaines River, IL. Chemosphere 2002, 49, 731–737. (23) Rayne, S.; Ikonomou, M. G.; Antcliffe, B. Rapidly increasing polybrominated diphenyl ether concentrations in the Columbia River system from 1992 to 2000. Environ. Sci. Technol. 2003, 37, 2847–2854. (24) Montgomery, D. C. Design and analysis of experiments, 3rd ed.; John Wiley & Sons: New York, 1991. (25) Stapleton, H. M.; Letcher, R. J.; Baker, J. E. Debromination of polybrominated diphenyl ether congeners BDE 99 and BDE 183 in the intestinal tract of the common carp (Cyprinus carpio). Environ. Sci. Technol. 2004, 38, 1054–1061. (26) Isosaari, P.; Lundebye, A. K.; Ritchie, G.; Lie, Ø.; Kiviranta, H.; Vartiainen, T. Dietary accumulation efficiencies and biotransformation of polybrominated diphenyl ethers in farmed Atlantic salmon (Salmo salar) Food Add. Contam. 2005, 22, 829-837.. (27) Kahnle, A. ; Hattala, K.; Stegemann, E. The true bass of New York, 2007; available athttp://www.dec.ny.gov/animals/7018.html. (28) Lepom, P.; Karasyova, T.; Sawal, G. Occurrence of polybrominated diphenyl ethers in freshwater fish from Germany. Organohalogen Compd. 2002, 58, 209–212. (29) Haglund, P. S.; Zook, D. R.; Buser, H.; Hu, J. Identification and quantification of polybrominated diphenyl dthers and methoxypolybrominated diphenyl ethers in Baltic Biota. Environ. Sci. Technol. 1997, 31, 3281–3287. (30) Vives, I.; Grimalt, J. O.; Lacorte, S.; Guillamo´n, M.; Barcelo´, D. Polybromodiphenyl ether flame retardants in fish from lakes in European high mountains and Greenland. Environ. Sci. Technol. 2004, 38, 2338–2344. (31) Olsson, A.; Valters, D.; Burreau, S. Concentrations of organochlorine substances in relation to fish size and trophic position: A study on perch (Perca fluviatilis L.). Environ. Sci. Technol. 2000, 34, 4878–4886.





Suggest Documents