Relative Potencies of Individual Polychlorinated Naphthalenes and

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Environ. Sci. Technol. 2000, 34, 3153-3158

Relative Potencies of Individual Polychlorinated Naphthalenes and Halowax Mixtures To Induce Ah Receptor-Mediated Responses A L A N L . B L A N K E N S H I P , * ,†,‡,§ K U R U N T H A C H A L A M K A N N A N , †,‡ S E R G I O A . V I L L A L O B O S , †,‡ D A N I E L L . V I L L E N E U V E , †,‡,| JERZY FALANDYSZ,⊥ TAKASHI IMAGAWA,∇ EVA JAKOBSSON,+ AND J O H N P . G I E S Y †,‡,| National Food Safety and Toxicology Center, Michigan State University, East Lansing, Michigan 48824, Institute for Environmental Toxicology, Michigan State University, East Lansing, Michigan 48824, ENTRIX, East Lansing, Michigan 48823, Zoology Department, Michigan State University, East Lansing, Michigan 48824, University of Gdan ˜ sk, PL 80-952 Gdan ˜ sk, Poland, National Institute for Resources and Environment, 16-3 Onogawa, Tsukuba, Ibaraki 305, Japan, and Wallenberg Laboratory, Department of Environmental Chemistry, Stockholm University, S-106 91 Stockholm, Sweden

Polychlorinated naphthalenes (PCNs) are ubiquitous environmental pollutants that are structurally similar to other polychlorinated diaromatic hydrocarbons (PCDHs), such as polychlorinated dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs), and biphenyls (PCBs). Despite being ubiquitous, much less is known about the fate, transport, and biological effects of individual PCN congeners than other PCDHs. The purpose of the current study was to utilize an in vitro assay (H4IIE-luc) to determine potencies relative to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) for 20 individual PCNs (from 75 possible congeners) and 6 Halowax mixtures. H4IIE rat hepatoma cells (H4IIE-Luc), which are stably transfected with an Ah receptor (AhR)-controlled luciferase reporter gene construct, respond specifically to AhR agonists and are thus a reasonable measure of AhRmediated, or dioxin-like, activity. The most potent congeners were 1,2,3,4,6,7-hexa-CN (PCN 66), 1,2,3,5,6,7hexa-CN (PCN 67), and 1,2,3,4,5,6,7-hepta-CN (PCN 73), with relative potencies as compared to TCDD of 0.004, 0.001, and 0.001, respectively. Significant structure-activity relationships were observed. For example, lateral substitution is an important determinant of AhR-mediated activity, but not sufficient, as illustrated by the inactivity of 2,3,6,7tetra-CN to elicit AhR-mediated activity. Relative potencies of the Halowax mixtures with AhR-mediated activity were 0.0089, 0.000038, and 0.0000018 for 1051, 1014, and 1013, respectively. The relative potencies derived from this study were applied to literature-derived data on concentrations of PCN congeners in environmental mixtures to assess the potential contribution of PCNs to total TCDD equivalents (TEQs) in environmentally weathered complex mixtures.

Introduction Polychlorinated naphthalenes (PCNs) are ubiquitous environmental pollutants that are members of the polychlorinated 10.1021/es9914339 CCC: $19.00 Published on Web 06/23/2000

 2000 American Chemical Society

diaromatic hydrocarbon (PCDH) class of compounds and are structurally similar to polychlorinated dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs), and biphenyls (PCBs). There are 75 possible PCN congeners with unique combinations of numbers and positions of chlorines (1, 2). PCNs were produced as mixtures (e.g., Halowaxes, Nibren waxes, Seekay waxes, and Clonacire waxes) for commercial applications such as protective coating materials, dielectric fluids, flame retardants, and even as fungicides (3, 4). In addition, PCNs were also present as byproducts in PCB formulations at levels up to 870 µg/g (5, 6). Recently, de novo synthesis mechanisms for PCN formation and emission from municipal waste incinerators have been reported (7). As with other PCDHs, the major mechanism of action for the toxicity of PCNs is related to their ability to bind to and activate the aryl hydrocarbon receptor (AhR), which is a cytosolic, ligand-activated transcription factor (Figure 1) (8, 9). The most characterized pathway involves translocation of the activated, cytosolic AhR to the nucleus where it binds with the AhR nuclear translocator protein (ARNT) to form a heterodimer, which subsequently binds to specific regions of DNA. Binding of the heterodimer results in modulation of gene transcription of genes that contain a dioxin responsive element (DRE). An additional AhR-dependent pathway has also been characterized involving protein phosphorylation which can occur in the absence of DNA binding of the AhR: ARNT complex which can lead to modulation of gene transcription (10-12). Toxic effects mediated through the AhR are species-, sex-, and tissue-specific, and include among other pleiotropic effects a characteristic wasting syndrome, thymic atrophy, immunosuppression, liver enlargement and necrosis, hyperplasia, chloracne, numerous biochemical effects, carcinogenesis, teratogenesis, and death (8). Exposure to PCNs has long been known to be associated with chloracne and lethality in occupationally exposed men (3, 13-15). In most of these cases of chloracne and mortality, however, the possibility cannot be ruled out that other contaminants such as 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) or PCBs could have been present at sufficient concentrations to elicit these effects. Experimental exposures of animals to PCN mixtures have resulted in chloracne in humans (15), “X-disease” in cattle (16), induction of cytochrome P450 enzyme activities and mortalities in chickens and eider ducklings (17), P450 induction in immature male wistar rats (18, 19), three-spined stickleback (20), and rainbow trout sac fry (21). Despite being ubiquitous in the environment and continuing releases to the environment, little is known about the fate, transport, and biological effects of individual PCN congeners. This is due, in part, to a deficiency of standards of individual congeners and a lack of sensitive, congenerspecific analytical methods. Now that many standards have been synthesized and more advanced analytical methods have been developed, such as capillary gas chromatography and high-resolution mass spectrometry (HRGC/HRMS), PCN * Corresponding author telephone: (517)432-6333; fax: (517)4322310; e-mail: [email protected]. † National Food Safety and Toxicology Center, Michigan State University. ‡ Institute for Environmental Toxicology, Michigan State University. § ENTRIX. | Zoology Department, Michigan State University. ⊥ University of Gdan ˜ sk. ∇ National Institute for Resources and Environment. + Stockholm University. VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 1. Model depicting mechanisms of Ah receptor (AhR) activation in H4IIE-luc cells. Refer to the text for a full description. In brief, pathway 1 depicts an AhR agonist entering a cell and interacting with an intracellular AhR, which then binds to dioxin-responsive elements (DREs) in the promoter region of AhR-responsive genes. Pathway 2 depicts an additional pathway by which gene expression can be modulated through activation of a protein phosphorylation pathway. congeners have been detected in environmental samples from North America (22-26), Europe (26-30), and Japan (31). Recently, the composition of PCN congeners in technical Halowax mixtures has been reported (32). Interest exists in assessing the risk of potential adverse effects posed by AhR-active or dioxin-like compounds, including PCNs, which are present as complex environmental mixtures. Development of mechanism-based cell bioassays to detect specific classes of compounds has enhanced the ability to screen such mixtures for AhR mediated or dioxinlike activity. H4IIE-luciferase cells, which are stably transfected with an AhR-controlled luciferase reporter gene construct, respond specifically to AhR agonists. The purpose of the current study was to utilize the H4IIE-luciferase bioassay to determine relative potencies for 20 individual PCNs and 6 Halowax mixtures compared to TCDD as a standard reference compound. Furthermore, the relative potencies from this assay were applied to literature-derived data on environmental levels of PCN congeners in Halowaxes and environmental mixtures to assess the potential contribution of PCNs to the AhR-mediated activity of a mixture expressed as total TCDD equivalents (TEQs).

Materials and Methods Chemicals. Halowax standards and 1,4-di-CN were obtained in methanol from Accustandard Inc. (New Haven, CT). Halowaxes 1000, 1001, and 1099 were 95% pure whereas Halowaxes 1014, 1051, and 1013 were 99% pure. 1,4-di-CN was 98.3% pure. 2-mono-CN was obtained from ICN Chemicals, Inc. (Irvine, CA) and was 99% pure. PCN congeners 1,2,4,5,6,8-hexa-CN, a mixture of 1,2,3,4,6,7 and 1,2,3,5,6,7-hexa-CN, 1,2,3,4,5,6,7-hepta-CN, and octa-CN were synthesized and purified at the Wallenberg Laboratory, Stockholm University, Sweden (33, 34). PCN congeners 1,2,6,8 and 1,2,4,6-tetra-CNs were synthesized at the National Institute for Resources and Environment, Tsukuba, Japan. Congener 1,2,6,8-tetra-CN was 93% pure and had 1,2,3,6,8and 1,2,4,5,6-penta-CNs and an unidentified hexa-CN as impurities. Congener 1,2,4,6-tetra-CN was 98% pure with 3154

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1,2,4,6,8-penta-CN and two tri-CNs as impurities. Other PCN congeners were obtained from Promochem (GmbH, Wesel, Germany) and were of >99% purity. Standards were analyzed for the presence of 2,3,7,8tetrasubtituted PCDDs or PCDFs and in the case of individual PCN congeners for the presence of other congeners using selected ion monitoring by low-resolution gas chromatography-mass spectrometry (GC-MS). PCDD and PCDF congeners were not detected at 100 pg/g level in the standards examined. At this level, the maximum possible response that could be expected to be contributed by contaminants is 600fold below the detection limit of the bioassay. Maximum concentrations tested in the bioassay varied for each congener and were limited by the mass of individual congeners that were available. Standards were prepared in isooctane prior to dosing cells. Cell Culture. Rat hepatoma cells stably transfected with an AhR-controlled luciferase reporter gene construct (H4IIEluc) were cultured in Dulbecco’s Modified Eagle’s Medium (DMEM) supplemented with 10% defined fetal bovine serum (FBS; Hyclone, Logan, UT) at 37 °C and 5% CO2 in a humidified atmosphere (35). For the bioassay, 96-well culture ViewPlates (Packard Instruments, Meriden, CT) were seeded with 250 µL of cell suspension at a density of 6 × 104 cells/ mL. Treatment of Cells. After 24 h, the medium was changed to a medium containing 10% charcoal-stripped FBS (Hyclone, Logan, UT), and the cells were dosed with either no treatment (blanks), solvent only, TCDD, or the various samples to be tested in a volume of 1.25 µL. Cells were dosed in triplicate with TCDD in isooctane (0.1-30 pg/well) or test agents dissolved in isooctane (a minimum of 6 concentrations were tested). Each experiment was repeated two or three times. Solvent controls for isooctane were not significantly different from blanks. Cells were evaluated microscopically for cellular degeneration and obvious toxicity. Since there was no direct toxicity, the lack of response in the H4IIE-luc cells to some treatments was not due to being near a toxic concentration. Luciferase activity were measured 3 days after dosing.

Luciferase activity was measured by incubating cells with LucLite reagent (Packard Instruments) for 20 min at room temperature. Light production, a measure of luciferase activity, was determined with a Dynatech ML 3000 luminometer at 30 °C. Data Analysis. If a complete dose-response curve was obtained, the data were converted to probit values and the concentration of PCN congener or Halowax mixture producing a response equivalent to 50% of the maximal response (EC50) produced by the standard (e.g., TCDD) was calculated in order to determine relative potencies. In cases in which incomplete dose-response curves were obtained, the data were treated similarly except that the concentration producing a response equivalent to 20% of the maximal response produced by the standard (e.g., TCDD) was calculated and used to identify PCN congeners with limited AhR-mediated activity and to estimate relative potency. Statistical significance between treatments and controls was determined by performing a Student’s t-test (p e 0.05). The coefficient of variation for relative potency calculations for samples that were tested in three independent experiments was between 10 and 15%. The relative potencies reported are the averages of two or three independent experiments in which there were three replicates per treatment. Calculation of TCDD Equivalents (TEQs). TEQs can be calculated for any sample for which there are both congenerspecific concentrations and a relative potency (REP), which is specific to an experiment, or a TCDD equivalency factor (TEF), which is consensus-derived from many relative potency values (36). TEQs contributed from PCN congeners were calculated as

FIGURE 2. Structures and relative potencies of the AhR-active PCN congeners analyzed in this study. The numbers in parentheses indicate the potency of each individual congener relative to 2,3,7,8tetrachlorodibenzo-p-dioxin (TCDD). Also shown are structures for TCDD and some of the inactive PCN congeners.

where n is any PCN congener.

Calculation of TEQs Due to PCNs in Environmental Samples. The TCDD relative potency factors (REPs) determined in this study were used to determine the potential contributions of PCNs to TEQs in various environmental samples (Table 2). This table is not meant to be all-inclusive of TEQ contributions due to PCNs because there are other PCN congeners that are present in environmental samples for which REPs are not available. Note that in contaminated sediments, PCNs can contribute most of the calculated TEQs in a sample. However, in most biological samples, the contribution of PCNs to total TEQs is approximately 1% (when considering the contributions from PCDDs, PCDFs, and PCBs). At some contaminated sites, the contributions of PCNs to the total TEQs can be up to 8% in biological samples and up to 58% in sediment samples (27, 37).

Results

Discussion

Individual PCN Congeners. Individual PCN congeners were tested for both their ability to produce a statistically significant response over solvent control and their ability to elicit a complete dose-response curve. Significant structure-activity relationships were observed. In general, full dose-response curves were obtained for more chlorinated congeners (e.g., penta-, hexa-, and hepta-CNs) whereas most of the lesser chlorinated congeners as well as octa-CN were inactive. Structures of TCDD and some of the AhR-active and inactive PCN congeners are shown along with their associated relative potencies (Figure 2). Representative dose-response curves for some of the hexa- and hepta-CNs are presented (Figure 3). Note that the term “inactive” is specific to the conditions in this assay. In other words, there was no significant AhRmediated activity up to the maximum concentration tested. Results for all congeners tested are summarized along with the greatest concentration tested and their maximum induction of luciferase normalized to the maximum elicited by TCDD (Table 1). None of the PCNs tested were able to elicit the same maximal activity or efficacy as TCDD. Halowax Mixtures. Six halowaxes were tested for their ability to induce luciferase activity in H4IIE-luc cells. Three of the halowaxes were capable of inducing luciferase activity and the other three were inactive (Table 1). Halowaxes 1000, 1001, and 1099, which were found to be inactive at the concentrations used in this study, are composed of mostly mono-, di-, tri-, and tetra-CNs (14). However, Halowaxes 1013, 1014, and 1051, which were found to be active, are composed of mostly tetra- through octa-CNs (14).

As a chemical class, relatively little is known about the environmental fate or toxicity of PCNs as compared to PCDDs, PCDFs, and PCBs. In this current investigation, relative potencies (REPs) were determined for the first time for individual PCN congeners using H4IIE-luc cells. Several of the PCN congeners proved to be AhR-active, as determined by their ability to induce luciferase activity through an AhRdependent pathway. REPs of the most potent PCN congeners compared to TCDD were in the range of 0.001-0.004. As a comparison, mammalian TEFs of some of the mono-orthoPCBs are in the range of 0.0001-0.0005. The most toxic nonortho-PCBs, such as 3,3′,4,4′,5-pentachlorobiphenyl (PCB 126), 3,3′4,4′,5,5′-hexachlorobiphenyl (PCB 169), and 3,3′,4,4′tetrachlorobiphenyl (PCB 77), have mammalian TEFs of 0.1, 0.01, and 0.0001, respectively (36). Thus, the REPs of some PCN congeners are in the same range as some PCB congeners. The REPs for hexa- and hepta-CNs indicate that the position of chlorine is an important determinant of activity mediated by the AhR. For example, shifting the position of a single chlorine on 1,2,3,4,5,6,7-hepta-CN (PCN 73) to become 1,2,3,4,5,6,8-hepta-CN (PCN 74) or the addition of a single chlorine to 1,2,3,4,5,6,7-hepta-CN (PCN 73) to become 1,2,3,4,5,6,7,8-octa-CN (PCN 75) eliminates all AhR-mediated activity. With other classes of halogenated aromatic hydrocarbons, TEFs generally correlate with the presence of lateral substitution and the absence of nonlateral substitution. As illustrated by the inactivity of 2,3,6,7-tetra-CN, lateral substitution with PCNs is important but not sufficient to cause AhR-mediated activity.

ng of TEQ/kg of sample ) n

∑[total concn of PCN n in sample (ng/kg) × i)1

REP for PCN n] (1)

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FIGURE 3. Dose-dependent induction of luciferase activity in the rat hepatoma H4IIE-luc cell bioassay by individual hexachlorinated (hexa-CN) and heptachlorinated (hepta-CN) naphthalene congeners as compared to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). The response is relative to a solvent control. Each point represents the mean of at least three replicates. The standard deviation for points within each independent experiment was under 10% of the mean.

TABLE 1. Summary of Relative Potencies and Efficacy for Individual PCN Congeners and Halowax Mixtures H4IIE-luc (this study) PCN no.

PCN structure

2 5 6 10 17 27 33 36 40 48 53 54 64 66 67 66/67 68 69 70 71 73 74 75

2-CN 1,4-di-CN 1,5-di-CN 2,3-di-CN 1,2,7-tri-CN 1,2,3,4-tetra-CN 1,2,4,6-tetra-CN 1,2,5,6-tetra-CN 1,2,6,8-tetra-CN 2,3,6,7-tetra-CN 1,2,3,5,8-penta-CN 1,2,3,6,7-penta-CN 1,2,3,4,5,7-hexa-CN 1,2,3,4,6,7-hexa-CN 1,2,3,5,6,7-hexa-CN 50:50 mix of 66/67 1,2,3,5,6,8-hexa-CN 1,2,3,5,7,8-hexa-CN 1,2,3,6,7,8-hexa-CN 1,2,4,5,6,8-hexa-CN 1,2,3,4,5,6,7-hepta-CN 1,2,3,4,5,6,8-hepta-CN 1,2,3,4,5,6,7,8-octa-CN

1000 1001 1013 1014 1051 1099

relative potency at

EC50a

NQ NQ* NQ NQ NQ NQ NQ NQ NQ* NQ NQ 1.7E-04 NA 4.0E-03 1.0E-03 1.3E-03 1.5E-04 NA 5.9E-04 NQ 1.0E-03 NQ NQ Halowax Mixtures NQ NQ 1.8E-06 3.8E-05 8.9E-03 NQ

% of TCDD maxb

tested (ng/well)c

H4IIE ERODd

7 24 8 7 19 8 18 8 35 14 7 57

1250 1250 625 625 625 2143 1025 1.875 144 12.5 12.5 12.5

85 65 75 67

12.5 12.5 155 12.5

83 9 70 3 8

12.5 127.5 128.75 12.5 2.5

NA NA NA NA NA NA NA NA NA NA NA NA 2.0E-05 NA 2.0E-03 NA 2.0E-03 2.0E-03 NA 7.0E-06 3.0E-03 NA NA

15 14 54 69 71 6

12.5 12.5 1250 1250 1250 1250

NA NA NA NA NA NA

a NQ indicates that the relative potency was below the ability of this bioassay to quantify. Congener data marked with an asterisk (*) exhibited slight activity; however, a relative potency could not be calculated because a full dose-response curve was not obtained. NA indicates that the congener was not analyzed in the bioassay. b Percentage of the maximum response induced by TCDD. c Maximum concentrations tested in the bioassay varied for each congener and were limited by the mass of individual congeners that were available. d Data from ref 46. The identification of these congeners were not fully confirmed. Also, the authors noted that the calculated relative potency for no. 71 was highly uncertain.

Interestingly, studies with polybrominated naphthalenes (PBNs) have reported that PBNs are more potent than PCNs (44, 45). For example, the LD50 for 2,3,6,7-tetra-BN in guinea pigs is approximately 88 times less potent than 2,3,7,8-TCDD, whereas a LD50 dose was not reached for the corresponding 3156

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2,3,6,7-tetra-CN congener (it is listed as greater than 2000 times less potent than 2,3,7,8-TCDD) (44). The difference in potency between the 2,3,6,7-tetra-substituted PBN and PCN, which are both approximate isosteres of 2,3,7,8-TCDD with a planar configuration, is thought to be due to the distance

TABLE 2. TEQs Contributed by PCNs, PCBs, and PCDD/Fs in Environmental Samples TEQs contributed by chemical classa sample

PCNsb

PCBs

PCDD/Fs

total TEQs

ref

pike muscle guillemot egg sediment urban air human adipose (range of 76 samples) human milk

117 pg/g, lipid 75 pg/g, lipid 5.16 ng/g, dry wt 0.4 fg/m3 0.1-3 pg/g, lipid 0.12-0.55 pg/g, lipid

674 pg/g, lipid 5012 pg/g, lipid 2.99 ng/g, 6.443 fg/m3 18.5-55.6 pg/g, lipid 20-56 pg/g, lipid

625 pg/g, lipid 1976 pg/g, lipid 0.66 ng/g, dry wt NA 10.2-228.9 pg/g, lipid 13-43 pg/g, lipid

1416 pg/g, lipid 7063 pg/g, lipid 8.81 ng/g, dry wt 6.843 fg/m3 28.8-287.5 pg/g, lipid 33-100 pg/g, lipid

27 27 37-39 40 23, 41, 42 43

a

Actual concentrations of PCNs, PCBs, and PCDD/Fs can be found in the cited references. TEQs for PCBs, PCDDs, and PCDFs were calculated on a congener-specific basis using the WHO TEFs (36). b TEQs were calculated for PCN congeners 66/67 (mix) and 73 by using relative potencies from this present work; for congener 69, relative potency was 0.002 (46).

between the lateral halogens. The distance between lateral halogens is 8.06 Å for the 2,3,6,7-tetra-BN and 7.85 Å for the 2,3,6,7-tetra-CN as compared to 10 Å for 2,3,7,8-TCDD. In addition, bromine is more polarizable than chlorine, and the additional electrostatic attraction could modulate the potency. Thus, it is possible that the relative inactivity of the 2,3,6,7-tetra-CN in this current study and in previous studies is due to a poor fit into the AhR as compared to the 2,3,6,7tetraBN. As more molecular descriptors become available for individual PCNs and PBNs, key features may be identified to provide a better understanding of potential structureactivity relationships. The results from this study compare well to previous studies on PCNs for 1,2,3,5,6,7-hexa-CN (PCN 67) and 1,2,3,4,5,6,7-hepta-CN (PCN 73) but are considerably different for other congeners (Table 1) (18, 19, 46). For example, Campbell et al. (18, 19) reported that octa-CN (PCN 75) and 1,2,3,4,5,6,8-hepta-CN (PCN 74) induced microsomal AHH activity in male rats, but both of these congeners were inactive in the current study. A possible explanation for differences between the present study and previous studies is that other PCDHs were present as impurities at greater concentrations in previous studies as compared to the present study (47). Also, some of the previous studies indicated that identification of congeners was not fully confirmed. Other reasons that might account for differences include different methods used to calculate relative potencies and possible differences between relative potencies determined in in vivo and in vitro toxicity models. The REPs reported here are relative potencies, not TEFs. REPs are usually derived from individual experiments using a single end point, whereas TEFs are consensus values that are derived from many different relative potencies for a variety of end points (36). The REPs derived in this study may not be appropriate for application to birds and fish, as evidenced by the different PCB TEFs for mammals, birds, and fish developed by the World Health Organization (36). For example, Engwall et al. (17) observed EROD induction and lethality in chicken embryos with exposure to a mixture of 1,2,3,4,6,7/1,2,3,5,6,7-hexa-CNs 66 and 67 (ED50 for EROD induction ) 0.06 mg/kg of egg), but when they tested 1,2,3,4,5,6,7-hepta-CN (PCN 73), they observed less EROD induction (the greatest dose, 3 mg/kg, caused only 50% of the maximal induction caused by the hexa-CN mix). Similarly, REPs of Halowax mixtures determined in H4IIE-luc cells (Halowax 1051 > 1014 > 1013) were different from REPs determined in medaka fish (Halowax 1014 > 1013 > 1051) (48). Interestingly, the relative rank order potency of CYPIA expression in rainbow trout sac fry is consistent with our results with the order being 66/67 > 73 > Halowax 1014 (21). Discrepancies between mammalian (H4IIE cells), fish, and avian (chicken egg)-derived REPs also occurs with PCBs as illustrated by the WHO TEFs (36) of 0.01, 0.00005, and 0.001 for 3,4,5,3′,4′,5′-hexachlorobiphenyl (PCB 169) for mammals,

fish, and birds, respectively. Thus, there can be significant species-specific differences that must be considered before relative potencies (or TEFs) are applied to environmental data. In addition, the REPs, like TEFs, do not account for differences in toxicokinetics among congeners. When assessing the toxicity of complex mixtures, a decision is usually made either to use a congener-specific approach or a surrogate complex mixture (e.g., Aroclors for PCBs, Halowaxes for PCNs, etc.). In the case of PCBs, toxicity is better correlated to TEQs (calculated from a congenerspecific approach) rather than to total PCBs for both mink (49) and piscivorous birds (50). Until a complete congenerspecific analysis of Halowax mixtures is available, as it is for Aroclors, it is unclear if TEQs or total PCN concentrations better predict toxicity. However, it is apparent from the toxicity of halowaxes and homologue groups of PCNs (such as pentachlorinated and hexachlorinated naphthalenes) that the spectrum of adverse effects is similar to those elicited by exposure to TCDD (51). Few studies have attempted to account for the sources of bioassay-derived TCDD-equivalents (TCDD-EQ) in environmental samples through a mass balance approach using a combination of both instrumental and bioassay analysis (24). In some of these studies, the bioassay detected more TCDD-EQ than could be accounted for by the TEQs predicted from instrumental analysis alone (including PCBs, PCDDs, and PCDFs). This suggests that additional classes of chemicals were contributing to the TCDD-EQ (24, 52). From a survey of the literature, PCNs generally contribute 1% or less of TEQs (on a tissue residue basis) in biological samples. However, in areas contaminated with PCNs, they can contribute substantially to the TEQs (27, 37). It is also possible that significant nonadditive interactions may be observed with PCNs as has been shown for PCBs (35). In laboratory studies with mixtures containing PCDDs, PCDFs, and PCBs, AhR antagonist activities of certain PCB congeners, including the major environmental contaminant PCB 153, have been reported in several experimental systems (36). While in vitro bioassays are useful to initially rank PCN congeners, such as was done here with the H4IIE-luc cells to predict the toxicological significance of PCNs, additional information will be necessary. This information includes (i) uptake from various environmental matrixes and (ii) toxicokinetics in an in vivo situation. For example, the bioavailability of PCNs may be limited in sediments that are rich in organic carbon because PCNs bind strongly to sediment organic carbon (37). Absorption, body distribution, and metabolism are important factors that contribute to the in vivo potency of an individual PCDD or PCDF (53). Similar effects are likely for PCNs. For example, the three most toxic PCNs (66, 67, and 73) are also the congeners most likely to biomagnify (54, 55). However, for other congeners there is no apparent correlation between biomagnification potential and REP (15, 30). VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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In conclusion, potencies relative to TCDD were determined for several individual PCN congeners for the first time in the H4IIE-luc bioassay. Several congeners were potent and dose-dependent inducers of luciferase activity. From the data presented here, it should be possible to ascertain the relative contribution of PCN to the total TEQ in environmental samples. As more congeners become available in a purified form, it will be possible to develop a comprehensive list of relative potencies for all of the active PCN congeners that are ubiquitous environmental pollutants.

Acknowledgments We gratefully acknowledge Dr. Jac Aarts for the generous gift of the cell line used in this study. We would also like to thank Sarah Cholger and Minghua Nie for their assistance in preparation of samples, Emily Nitsch for her assistance in the cell culture facilities, and the rest of the Aquatic Toxicology Laboratory for their support. This work was supported in part by a grant from the Chlorine Chemistry Council of the Chemical Manufacturers Association, the NIEHS Superfund Basic Research (Grant NIH-ES-04911), the U.S. EPA Biology Panel of the Exploratory Research Program (R82537-01), and a cooperative agreement between the U.S. EPA Office of Water and Michigan State University (CR822983-01-0). A.L.B. was supported by NIEHS Postdoctoral Training Grant ES04911.

(25) (26) (27) (28) (29) (30) (31) (32) (33) (34) (35) (36)

Literature Cited (1) Wiedmann, T.; Ballschmiter, K. Fresenius J. Anal. Chem. 1993, 346, 800-804. (2) Nakano, T.; Fujimori, K.; Takaishi, Y.; Umeda, H. Rep. Hyogo Prefect. Inst. Environ. Sci. 1993, 25, 33-40 (in Japanese with abstract in English). (3) Kover, F. Environmental hazard assessment of chlorinated naphthalenes; EPA 560/8-75-001; U.S. Environmental Protection Agency: Washington, DC, 1975. (4) Falandysz, J. Environ. Pollut. 1998, 101, 77-90. (5) Vos, J. G.; Koeman, J. H.; van der Maas, H. L.; ten Noever de Brauw, M. C.; de Vos, R. H. Food Cosmet. Toxicol. 1970, 8, 625635. (6) Haglund, P.; Jakobsson, E.; Asplund, L.; Athanasiadou, M.; Bergman, A. J. Chromatogr. 1993, 634, 79-86. (7) Iino, F.; Imagawa, T.; Takeuchi, M.; Sadakata, M. Environ. Sci. Technol. 1999, 33, 1038-1043. (8) Poland, A.; Knutson, J. C. Annu. Rev. Pharmacol. Toxicol. 1982, 22, 517-554. (9) Gasiewicz, T. A. In Handbook of Pesticide Toxicology; Hayes, W. J., Laws, E. R., Eds.; Academic Press: San Diego, 1991; pp 11911270. (10) Blankenship, A. Studies on the mechanism of protein kinase activation by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Ph.D. Thesis, University of California, Davis, CA, 1994. (11) Blankenship, A.; Matsumura, F. Mol. Pharmacol. 1997, 52, 667675. (12) Blankenship, A.; Matsumura, F. Environ. Toxicol. Pharmacol. 1997, 3, 211-220. (13) Greenburg, L.; Mayers, M. R.; Smith, A. R. J. Ind. Hyg. Toxicol. 1939, 21, 29-39. (14) Brinkman, U. A.; Reymer, H. G. M. J. Chromatogr. 1976, 127, 203-243. (15) Hayward, D. Environ. Res. 1998, 76, 1-18. (16) Bell, W. B. J. Am. Vet. Med. Assoc. 1954, 124, 289-90. (17) Engwall, M.; Brunstrom, B.; Jakobsson, E. Arch. Toxicol. 1994, 68, 37-42. (18) Campbell, M. A.; Bandiera, S.; Robertson, L.; Parkinson, A.; Safe, S. Toxicology 1981, 22, 123-132. (19) Campbell, M. A.; Bandiera, S.; Robertson, L.; Parkinson, A.; Safe, S. Toxicology 1983, 26, 193-205. (20) Holm, G.; Norrgren, L.; Andersson, T.; Thuren, A. Aquat. Toxicol. 1993, 27, 33-50. (21) Pesonen, M.; Teivainen, P.; Lundstrom, J.; Jakobsson, E.; Norrgren, L. Arch. Environ. Contam. Toxicol. 2000, 38, 52-58. (22) Furlong, E. T.; Carter, D. S.; Hites, R. A. J. Great Lakes Res. 1988, 14, 489-501. (23) Williams, D. T.; Kennedy, B.; LeBel, G. L. Chemosphere 1993, 27, 795-806. (24) Giesy, J. P.; Jude, D. J.; Tillitt, D. E.; Gale, R. W.; Meadows, J. C.; Zajieck, J. L.; Peterman, P. H.; Verbrugge, D. A.; Sanderson, J. 3158

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 15, 2000

(37) (38) (39) (40) (41) (42) (43) (44) (45)

(46) (47) (48) (49) (50) (51) (52) (53) (54) (55)

T.; Schwartz, T. R.; Tuchman, M. L. Environ. Toxicol. Chem. 1997, 16, 713-724. Kannan, K.; Yamashita, N.; Imagawa, T.; DeCoen, W.; Khim, J. S.; Day, R. M.; Summer, C. L.; Giesy, J. P. Environ. Sci. Technol. 2000, 34, 566-572. Harner, T.; Kylin, H.; Bidleman, T. F.; Halsall, C.; Strachan, W. M. J. Environ. Sci. Technol. 1998, 32, 3257-3265. Jarnberg, U.; Asplund, L.; Dewit, C.; Grafstrom, A.-K.; Haglund, P.; Jansson, B.; Lexen, K.; Strandell, M.; Olsson, M.; Jonsson, B. Environ. Sci. Technol. 1993, 27, 1364-1374. Jarnberg, U. G.; Asplund, L. T.; Egeback, A. L.; Jansson, B.; Unger, M.; Wideqvist, U. Environ. Sci. Technol. 1999, 33, 1-6. Falandysz, J.; Strandberg, L.; Bergqvist, P. A.; Kulp, S. E.; Strandberg, B.; Rappe, C. Environ. Sci. Technol. 1996, 30, 32663274. Falandysz, J.; Rappe, C. Environ. Sci. Technol. 1996, 30, 33623370. Imagawa, T.; Yamashita, N. J. Environ. Chem. 1996, 6, 495-501 (in Japanese). Falandysz, J.; Kawano, M.; Ueda, M.; Matsuda, M.; Kannan, K.; Giesy, J. P.; Wakimoto, T. J. Environ. Sci. Health Part A 2000, 35, 281-298. Jakobsson, E.; Eriksson, L.; Bergman, A. Acta Chem. Scand. 1992, 46, 527-532. Jakobsson, E.; Lonnberg, C.; Eriksson, L. Acta Chem. Scand. 1994, 48, 891-898. Sanderson, J. T.; Aarts, J. M. M. J. G.; Brouwer, A.; Froese, K. L.; Denison, M. S.; Giesy, J. P. Toxicol. Appl. Pharmacol. 1996, 137, 316-325. Van den Berg, M.; Birnbaum, L.; Bosveld, B. T. C.; Brunstrom, B.; Cook, P.; Feeley, M.; Giesy, J. P.; Hanberg, A.; Hasegawa, R.; Kennedy, S. W.; Kubiak, T.; Larsen, J. C.; van Leeuwen, F. X. R.; Liem, A. K. D.; Nolt, C.; Peterson, R. E.; Poellinger, L.; Safe, S.; Schrenk, D.; Tillitt, D.; Tysklind, M.; Younes, M.; Waern, F.; Zacharewski, T. Environ. Health Perspect. 1998, 106, 775-792. Kannan, K.; Imagawa, T.; Blankenship, A. L.; Giesy, J. P. Environ. Sci. Technol. 1998, 32, 2507-2514. Kannan, K.; Watanabe, I.; Giesy, J. P. Toxicol. Environ. Chem. 1998, 67, 135-146. Kannan, K.; Maruya, K. A.; Tanabe, S. Environ. Sci. Technol. 1997, 31, 1483-1488. Harner, T.; Bidleman, T. Atmos. Environ. 1997, 31, 4009-4016. Williams, D. T.; LeBel, G. L. Chemosphere 1991, 22, 1019-1028. LeBel, G. L.; Williams, D. T.; Benoit, F. M.; Goddard, M. Chemosphere 1990, 21, 1465-1476. Lunden, A.; Noren, K. Arch. Environ. Contam. Toxicol. 1998, 34, 414-423. McConnell, E. E.; McKinney, J. D. Toxicol. Appl. Pharmacol. 1978, 45, 298. Goldstein, J. A. In Halogenated Biphenyls, Terphenyls, Naphthalenes, Dibenzodioxins, and Related Products; Kimbrough, R. D., Ed.; Elsevier/North-Holland Biomedical Press: Amsterdam, 1980; pp 151-190. Hanberg, A.; Waern, F.; Asplund, L.; Haglund, E.; Safe, S. Chemosphere 1990, 20, 1161-1164. Koistinen, J.; Sanderson, J. T.; Giesy, J. P.; Nevalainen, T.; Paasivirta, J. Environ. Toxicol. Chem. 1996, 15, 2028-2034. Villalobos, S. A.; Papoulias, D.; Meadows, J.; Blankenship, A. L.; Pastva, S. D.; Kannan, K.; Tillitt, D. E.; Giesy, J. P. Environ. Toxicol. Chem. 2000, 19, 432-440. Leonards, P. E. G.; de Vries, T. H.; Minnaard, W.; Stuijfzand, S.; de Voogt, P.; Cofino, W. P.; van Straalen, N. M.; van Hattum, B. Environ. Toxicol. Chem. 1995, 14, 639-652. Giesy, J. P.; Ludwig, J. P.; Tillitt, D. E. Environ. Sci. Technol. 1994, 28, 128-135. Kimbrough, R.; Buckley, J.; Fishbein, L.; Flamm, G.; Kasza, L.; Marcus, W.; Shibko, S.; Teske, R. Environ. Health Perspect. 1978, 24, 173-184. Jones, P. D.; Giesy, J. P.; Newsted, J. L.; Verbrugge, D. A.; Beaver, D. L.; Ankley, G. T.; Tillitt, D. E.; Lodge, K. B.; Niemi, G. J. Arch. Environ. Contam. Toxicol. 1993, 24, 345-354. Van den Berg, M.; De Jongh, J.; Poiger, H.; Olson, J. R. Crit. Rev. Toxicol. 1994, 24, 1-74. Asplund, L.; Jansson, B.; Sundstrom, G.; Brandt, I.; Brinkman, U. A. T. Chemosphere 1986, 15, 619-628. Asplund, L.; Jakobsson, E.; Haglund, P.; Bergman, A. Chemosphere 1994, 28, 2075-2086.

Received for review December 24, 1999. Revised manuscript received May 8, 2000. Accepted May 8, 2000. ES9914339