Chicken Embryo Hepatocyte Bioassay for Measuring Cytochrome

Jan 29, 1996 - Polychlorinated Biphenyls and Organochlorine Pesticides in Plasma and the Embryonic Development in Lake Erie Water Snakes (Nerodia sipe...
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Environ. Sci. Technol. 1996, 30, 706-715

Chicken Embryo Hepatocyte Bioassay for Measuring Cytochrome P4501A-Based 2,3,7,8-Tetrachlorodibenzop-dioxin Equivalent Concentrations in Environmental Samples SEAN W. KENNEDY,* ANGELA LORENZEN, AND ROSS J. NORSTROM Environment Canada, Canadian Wildlife Service, National Wildlife Research Centre, Hull, Quebec, Canada K1A 0H3

A bioassay that uses chicken embryo hepatocyte (CEH) primary cultures for measuring 2,3,7,8-tetrachlorodibenzo-p-dioxin equivalent (TCDD-EQ) concentrations in extracts prepared from wild bird eggs contaminated with complex mixtures of polychlorinated biphenyls (PCBs), dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs), and other halogenated aromatic hydrocarbons is described. The CEH bioassay uses an efficient method for measuring ethoxyresorufin-Odeethylase (EROD) in hepatocytes cultured in 48-well plates to obtain bioassay-derived TCDD-EQ concentrations (TCDD-EQbio). Induction equivalency factors (IEFs) were determined for 39 PCB congeners, 2,3,7,8TCDD, 1,2,3,7,8-PCDD, and 2,3,7,8-TCDF. TCDD-EQbio concentrations in eggs from herring gulls, Larus argentatus, and great blue herons, Ardea herodias, were highly correlated with TCDD-EQ concentrations calculated (TCDD-EQcalc) from chemical residue data and IEFs (r2 ) 0.977, slope ) 0.99). Unlike other in vitro bioassays, 2,2′,3,4,4′,5′-hexachlorobiphenyl (PCB 138) made a significant contribution to TCDD-EQcalc in CEH cultures. TCDD-EQbio concentrations were also highly correlated to TCDD-EQcalc obtained using toxic equivalency factors (TEFs) derived from various in vivo and in vitro toxic and biochemical end points. The CEH bioassay is a cost-effective method for measuring TCDD-EQbio concentrations in environmental samples.

Introduction Polychlorinated biphenyls (PCBs), dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs), and structurally related * Address correspondence to this author; telephone: (819) 9976077; fax: (819) 953-6612; e-mail address: [email protected].

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halogenated aromatic hydrocarbons (HAHs) are industrial compounds or byproducts that are widespread environmental contaminants found in almost every matrix imaginable, including soil, air, water, and human and wildlife tissues (1). Global contamination with HAHs is of concern because many of these compounds are toxic and they are relatively resistant to degradation. In experimental animals, effects include dermal toxicity, hepatotoxicity, teratogenesis, immunotoxicity, tumor promotion, hormonal and neurobehavioral changes, and lethality (2-6). HAHs or their metabolites have been associated with some of these effects in fish and other wildlife (7-11). Determination of the potential toxicity of environmental concentrations of HAHs is difficult because (1) there are hundreds of isomers and congeners with toxic potencies that range over several orders of magnitude, (2) environmental mixtures of HAHs are extremely complex, and (3) there are large interspecies differences in sensitivity to these chemicals. One approach that attempts to estimate the toxic potential of environmental mixtures of PCBs, PCDDs, and PCDFs uses the toxic equivalency (TEQ) concept (2, 12-14). This concept exploits the results of many studies with laboratory animals and cultured cells that indicate that some of the most toxic compounds (the 2,3,7,8substituted PCDDs and PCDFs) cause similar effects, but differ in potency. Several, but not all, of these effects appear to be mediated by the aryl hydrocarbon (Ah) receptor (2, 15-19). The most toxic compound is 2,3,7,8-TCDD, and estimates of the toxic potential of mixtures of PCDDs, PCDFs, and PCBs are expressed in units of TCDD equivalents (TCDD-EQs), that is, the concentration of 2,3,7,8TCDD that would be expected to produce the same type and degree of response. TCDD-EQ concentrations are determined by using two general approaches: they are either calculated or bioassayderived. Calculated TCDD-EQ (TCDD-EQcalc) is the sum of the products of the concentrations of PCBs, PCDDs, and PCDFs that elicit 2,3,7,8-TCDD-like toxic effects and their toxic equivalency factors (TEFs). The TEF of a compound is an estimate of its toxic potency relative to that of 2,3,7,8TCDD as determined from various in vivo and in vitro studies. TEFs are end point- and species-dependent, but internationally accepted TEFs for PCDDs and PCDFs that are based on a variety of end points have been used for human risk assessment purposes (20, 21). TEFs for some of the non- and mono-ortho-substituted PCBs have also been proposed (2, 12, 14). These TEFs occasionally have been applied to wildlife (22-25), although this use has seldom been validated for the species in question. There are several problems with using TEFs for risk assessments. Among these are the following: (1) the approach assumes that the combined toxic effects of the components of a mixture are additive, therefore neglecting possible synergism or antagonism; (2) mammalian-derived TEF indices may not be suitable for some toxic end points and for some species (26-28); (3) TEFs are established for only a limited number of compounds, and there may be compounds that are not routinely detected by analytical protocols for quantification that are left out of risk characterizations; and (4) cost. Methods for quantifying PCDDs, PCDFs, and some of the PCB congeners of greatest

0013-936X/96/0930-0706$12.00/0

 1996 American Chemical Society

concern (the non-ortho-substituted congeners) require extensive sample cleanup and gas chromatography-mass spectrometry (GC-MS) analysis. For several years there has been interest in the possibility of developing cost-effective bioassays to obtain bioassayderived TCDD-EQ (TCDD-EQbio) concentrations that integrate the biological effects of complex mixtures of PCBs, PCDDs, PCDFs, and other HAHs (12, 29-33). With this approach, the toxic or biochemical potency of a mixture of HAHs in a biological system (e.g., cultured cells) is compared with the potency of 2,3,7,8-TCDD in the same biological system. Most bioassay research and development has used rat hepatoma H4IIE cells (29, 31, 32), a system that is sensitive to cytochrome P4501A (CYP1A) induction, an Ah receptor-mediated response that is associated with many of the biochemical and toxic effects of PCBs, PCDDs, and PCDFs (2, 17). The toxic potencies of various PCDD, PCDF, and PCB congeners in rats correlate quite well with potencies to induce CYP1A catalytic activity in rat hepatoma H4IIE cells (2). CYP1A activity is usually estimated by using either the ethoxyresorufin-O-deethylase (EROD) or the aryl hydrocarbon hydroxylase (AHH) assay, and the EROD (or AHH) induction potency of individual compounds or complex mixtures of compounds in H4IIE cells can be used as a surrogate for toxic potency. Several studies have shown that H4IIE cells can be used to measure TCDD-EQbio concentrations in extracts prepared from contaminated environmental samples (7, 29, 31, 32, 34). Until recently, methods for measuring EROD (or AHH) activity in cell cultures were time-consuming, and rigorous validation of CYP1A-based bioassays for measuring TCDDEQ concentrations was difficult. Therefore, we developed methods for measuring EROD activity in cultured hepatocytes and cellular fractions that are much more efficient than previous methods (35-38). Primary chicken embryo hepatocyte (CEH) cultures were used for this research because we are attempting to develop avian cell culture bioassays that measure TCDD-EQs that are relevant to birds, and we are also investigating porphyrin accumulation as a biochemical marker of wild bird exposure to HAHs (9, 39). CEH primary cultures are useful for studying the effects of HAHs and other xenobiotics on both CYP1A induction and porphyrin accumulation (40-43). Therefore, we reasoned that it may be possible to develop a bioassay with CEH cultures that measures TCDD-EQbio concentrations that are based on two different biochemical end points (44). This paper describes the EROD component of the CEH bioassay and shows how it was used to measure TCDDEQbio concentrations in herring gull eggs collected from the Great Lakes and great blue heron eggs collected from the west coast of Canada. These samples represent different patterns of HAH contamination: gulls had high PCB contamination, but relatively low PCDD and PCDF levels (45) (Hebert et al., in preparation), whereas herons were exposed to relatively high levels of PCDDs, but not PCBs (46, 47). TCDD-EQbio and TCDD-EQcalc concentrations in the samples were compared.

Experimental Section Sources of HAH-Contaminated Eggs and Standards. Frozen (-20 °C) homogenates of herring gull (n ) 47) and great blue heron eggs (n ) 9) were obtained from the Specimen Bank at the National Wildlife Research Centre (Hull, Quebec, Canada). Herring gull eggs were collected

from various locations within the Great Lakes during the period of time from 1971 to 1992, and great blue heron eggs were collected from British Columbia in 1988 and 1990 (46, 47). 2,3,7,8-TCDD and 1,2,3,7,8-pentachlorodibenzo-pdioxin (PeCDD) were kindly provided by Drs. J. J. Ryan (Health Canada, Ottawa, Ontario, Canada) and A. T. C. Bosveld (U. Utrecht, The Netherlands), respectively. PCBs and 2,3,7,8-TCDF were from Ultra Scientific (Kingstown, RI) and were stated to be at least 99% pure by the supplier. PCB, PCDD, and TCDF standards were prepared in dimethyl sulfoxide (DMSO), and concentrations were confirmed by GC-MS or GC-ECD. The standards were not tested for the presence of trace levels of potent inducers of CYP1A. Extraction of HAHs and Preparation of Serial Dilutions. HAHs were extracted from egg homogenates using minor modifications of methods used in our laboratories for chemical residue analysis (48, 49), such that the final extracts added to the CEH bioassay contained all PCDDs, PCDFs, PCBs, structurally related nonpolar HAHs (e.g. polybrominated biphenyls, chloronaphthalenes, and diphenyl ethers), and chlorinated pesticides (e.g. DDE, Dieldrin, and Mirex). In brief, egg homogenates were dried with sodium sulfate, extracted with dichloromethane (DCM)/hexane (1: 1), and cleaned up by gel permeation chromatography followed by an alumina column. Extracts were transferred quantitatively from DCM/hexane into DMSO. The final DMSO solutions (referred to in the following as the extract stock solutions) were mixed by vortex, and serial dilutions were prepared in DMSO. Preparation and Dosing of CEH Cultures. Primary hepatocyte cultures were prepared from 19-day-old chicken embryos in 48-well plates, as described previously (35). After incubation for 24 h at 37 °C in an atmosphere of 95% air and 5% CO2, the medium was removed and replaced with fresh medium (0.5 mL). DMSO solutions (2.5 µL/ well) of individual compounds or egg extracts were added to the cultures. For each cell culture study, EROD doseresponse curves for TCDD were obtained in triplicate. Cells were incubated for another 24 h, the medium was removed, and plates were frozen on dry ice prior to transferring them to a freezer (-80 °C). For the studies described in this article, TCDD standards were administered at a maximal dose that was 1000 times higher than that required to obtain a reliable estimate of the EC50. The reason for using such a high dose was because we are also investigating the possibility of using porphyrin accumulation in CEH cultures as an end point that will provide additional information on the toxic potencies of extracts, which will complement the information provided by EROD induction potencies. As we reported previously (35, 37, 38), porphyrins accumulate at higher doses than those required to induce EROD. For studies where we require only the EROD component of the CEH bioassay, the highest dose of TCDD is 10 nM, thus reducing possible hazards when handling and disposing of this highly toxic chemical. EROD Assays. EROD assays were carried out in cell culture plates as described previously (35), and the reaction product (resorufin) was measured with a fluorescence plate reader (Cytofluor 2300, Millipore Ltd.). Separate 48-well plates were used for determining total protein concentration with the fluorescamine assay (36). EROD Dose-Response Curves. Data obtained with the fluorescence plate reader were fitted empirically to a modified (to include basal and maximal EROD activities as

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parameters) Gaussian curve:

y(d) ) Yb + (Ym - Yb) exp{-C[ln(d) - ln(dm)]2}

(1)

where

C)

ln(2) (ln(EC50) - ln(dm))2

where y(d) is the EROD activity at HAH concentration d, Yb is the basal EROD activity, Ym is the maximal EROD activity, and dm is the HAH concentration where EROD activity is maximal. The curve-fitting program (Sigma Plot, Jandel Scientific) provided the EC50 (defined as the lower concentration of HAH where EROD activity is midway between basal and maximal activity; Figure 1), Yb, Ym, dm, and standard errors for each of these parameters. Bioassay-Derived TCDD-EQs. TCDD-EQbio concentrations in eggs were obtained by using an equation similar to that used by Tillitt et al. (31) for rat hepatoma H4IIE cells:

TCDD-EQbio (ng/g) ) TCDD EC50 (ng/mL of medium) × extract EC50

(

(

)

)

volume of medium (mL) × volume of DMSO (µL) volume of extract stock (µL) (2) mass of sample (g)

(

)

Calculated TCDD-EQs. TCDD-EQcalc concentrations in herring gull and great blue heron eggs were obtained by summing the product of the concentrations of PCDDs, PCDFs, PCBs, and their respective TEFs or their EROD induction potencies relative to 2,3,7,8-TCDD (induction equivalency factors, IEFs) in CEH cultures:

TCDD-EQcalc )

∑(concentration

PCDD,PCDF,PCB × TEF(or IEF)PCDD,PCDF,PCB) (3)

The IEFs of PCB congeners, 1,2,3,7,8-PeCDD, and 2,3,7,8TCDF were obtained by dividing the EC50 of 2,3,7,8-TCDD by the EC50 of the compound. This ratio occasionally has been referred to as a TEF, but we prefer to use the term IEF because EROD induction is not a toxic response per se, it is a biochemical response that is associated with some of the toxic effects of PCBs, PCDDs, and PCDFs. In addition to CEH IEFs, two TEF schemes (Table 1) that are currently used for risk assessments were also used to obtain TCDD-EQcalc concentrations. (i) WHO/IPCS TEFs. The World Health OrganizationEuropean Centre for Environment and Health (WHO-ECEH) and the International Programme on Chemical Safety (IPCS) recently published interim TEFs for several PCBs (14). WHO/ IPCS TEFs for PCBs were used in conjunction with internationally accepted TEFs for PCDDs and PCDFs (20, 21) for what we refer to here as the WHO/IPCS scheme. (ii) International/Safe TEFs. For this commonly used scheme, internationally accepted TEFs for PCDDs and PCDFs (20, 21) were used in conjunction with Safe’s suggested TEFs for PCB congeners (12). Safe’s earlier TEFs for PCBs were used here rather than his more recently suggested (2) values because (a) they remain in relatively common use and it was worthwhile to compare TCDD-

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EQs obtained using these values with TCDD-EQs obtained using CEH-derived IEFs, and (b) the WHO/IPCS TEFs are similar to Safe’s recently suggested TEFs for most PCBs. The most noteworthy differences between Safe’s original TEF scheme and the WHO/IPCS TEFs scheme are for the nonortho-PCBs 77 and 169 (IUPAC nomenclature; see Table 1 for chemical structures) and several of the mono-orthosubstituted PCBs. Recent data suggest that these compounds may not be as toxic in mammals as originally indicated in short-term tests in rats (14). Chemical Residue Analysis. Aliquots (10 µL) of the stock solutions of each egg extract were partitioned from DMSO into isooctane (250 µL) and analyzed for PCBs (other than the non-ortho-substituted PCBs 37, 77, 81, 126, and 169) by GC-ECD (50). Addition of internal standards to the samples that were analyzed with the bioassay was not desirable because of possible effects on EROD induction. Therefore, concentrations of PCDDs, PCDFs, and the nonortho-PCBs 37, 77, 81, 126, and 169 were determined from separate analyses of the egg homogenates by GCMS. PCB 118 was used as an in situ internal standard to normalize the concentrations of the PCDDs, PCDFs, and nonortho-PCBs in the separate analyses to those in extracts that were added to the bioassay before calculation of TCDDEQs. PCB 118 was chosen for this purpose because it is a major congener and easily quantified by both GC-ECD and GC-MS. The ratio of PCB 118 concentrations in the two extractions was used to normalize the data from the separate analyses. Concentrations of PCDD/Fs and nonortho-PCBs were determined by using modifications of procedures described in Norstrom et al. (49) and Ford et al. (51) or by the method of Moisey (52). Residue data will be published elsewhere by Norstrom et al.

Results Dose-Response Curves. Dose-dependent effects of 2,3,7,8TCDD, 2,3,7,8-TCDF, and 11 PCB congeners on EROD activity are compared in Figure 1. For each curve, an increase in enzymatic activity at low doses was followed by a decrease in activity at higher doses. Other investigators have reported an increase in EROD activity at low doses of HAH exposure followed by a decrease in activity at higher doses in vivo and in cell cultures (summarized in ref 53). In some situations, decreased EROD activity at high doses is likely caused by cytotoxicity or competitive inhibition of enzymatic activity by HAHs, but Hahn et al. (53) have argued that cytotoxicity and/or enzyme inhibition may not provide a full explanation in all situations. Regardless of the mechanism(s) involved, dose-response data from all EROD inducers that we have studied in CEH cultures, including photooxidized products of tryptophan (54), can be fitted empirically to Gaussian curves. Parameters from fitted curves were used to estimate EC50’s and maximal responses. Induction Equivalency Factors of PCDDs, PCDF, and PCBs. 2,3,7,8-TCDF and 1,2,3,7,8-PeCDD gave doseresponse curves that were similar to the dose-response curve of 2,3,7,8-TCDD, and the EC50’s and maximal responses for these compounds were also similar to those of 2,3,7,8-TCDD (Figure 1, Table 1). IEFs for each compound were obtained by dividing the EC50 of 2,3,7,8-TCDD by the EC50 of the compound (Table 1). Thus 2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, and 2,3,7,8-TCDF were approximately equipotent, and the IEFs of PCB congeners ranged from 0 to 0.3. As found in several other in vitro and in vivo systems (reviewed in refs 2 and 12), the most potent PCB congeners

TABLE 1

EC50s, IEFs, and Percent Maximal EROD Activity Relative to Maximal Activity Elicited by 2,3,7,8-TCDD for PCDD/Fs and PCBs in CEH Bioassaya CEH bioassay EC50 (nM) mean (SE, n) 2,3,7,8-TCDD 1,2,3,7,8-PeCDD 2,3,7,8-TCDF

IEF

0.016 (0.004, 42) 1.0 0.014 (0.002, 3) 1.1 0.014 (0.003, 6) 1.1

% max acty 100 100 100

IEF Safe

H411E bioassay TEF IEF Tillitt IEF Clemons Int./Safe WHO/IPCS

1.0 0.01 0.09

1.0 NV 0.0064

1.0 1.1 0.03

1.0 0.5 0.1

1.0 0.5 0.1

3-CB 4-CB 3,3′-DiCB 3,4-DiCB 3,3′4-TriCB 3,4,4′-TriCB 3,3′,4,4′-TeCB 3,3′,4,5-TeCB 3,3′,4,5′-TeCB 3,3′,5,5′-TeCB 3,4,4′,5-TeCB 3,3′,4,4′,5-PeCB 3,3′,4,5,5′-PeCB 3,3′4,4′,5,5′-HxCB

PCB 2 PCB 3 PCB 11 PCB 12 PCB 35 PCB 37* PCB 77* PCB 78 PCB 79 PCB 80 PCB 81* PCB 126* PCB 127 PCB 169*

1500 (350, 3) no induction no induction 19 (1, 3) 22 (3, 3) 40 (3, 3) 0.51 (0.07, 20) 145 (26, 2) 4.1 (2.0, 2) 200 (100, 2) 0.094 (0.003, 3) 0.052 (0.004, 5) 3.4 (1.0, 5) 0.79 (0.08, 3)

Non-ortho-PCBs 0.00001 10 0.0 0 0.0 0 0.0008 70 0.0007 60 0.0004 40 0.03 70 0.0001 90 0.004 40 0.00008 5 0.2 100 0.3 80 0.005 10 0.02 60

NV NV NV NV NV NV 0.0009 NV NV NV 0.00004 0.3 NV 0.003

NV NV NV NV NV NV 0.00002 NV NV NV NV 0.02 NV 0.0005

NV NV NV NV NV NV 0.0008 NV NV NV 0.007 0.1 NV 0.001

NV NV NV NV NV NV 0.01 NV NV NV NV 0.1 NV 0.05

NV NV NV NV NV NV 0.0005 NV NV NV NV 0.1 NV 0.01

2,3′4,4′-TeCB 2,3′,4′,5-TeCB 2,3,3′4,4′-PeCB 2,3′,4,4′,5-PeCB 2′,3,3′,4,5-PeCB 2,3,3′,4,4′,5-HxCB 2,3,3′,4,4′,5′-HxCB 2,3′,4,4′,5,5′-HxCB

PCB 66* PCB 70* PCB 105* PCB 118* PCB 122 PCB 156* PCB 157* PCB 167

8.7 37 3.3 19.1 826 11.3 6.4 7.1

(2.7, 6) (5, 3) (1.5, 6) (7.1, 6) (53, 3) (2.1, 4) (0.2, 2) (1.2, 3)

Mono-ortho-PCBs 0.002 10 0.0004 10 0.005 10 0.001 30 0.00002 60 0.001 40 0.002 30 0.002 10

NV NV 0.0007 0.000009 NV 0.00009 0.00006 NV

NV NV 0.000008 0.000001 NV 0.00006 NV NV

NV NV 0.00003 0.00001 NV 0.00005 NV NV

NV NV 0.001 0.001 NV 0.001 0.001 0.001

NV NV 0.0001 0.0001 NV 0.0005 0.0005 0.00001

2,2′-DiCB 2,2′,4,4′-TeCB 2,2′,5,5′-TeCB 2,2′,4,5,5′-PeCB 2,3,3′,4′,6-PeCB 2,2′,3,3′,4,4′-HxCB 2,2′,3,4,4′,5′-HxCB 2,2′,4,4′,5,5′-HxCB 2,2′,3,3′,4,4′,5-HpCB 2,2′,3,4,4′,5,5′-HpCB 2,2′,3,3′,4,4′,5,5′-OcPCB

PCB 4 PCB 47* PCB 52* PCB 101* PCB 110* PCB 128* PCB 138* PCB 153* PCB 170* PCB 180* PCB 194*

no no no no 300 13 13.2 no 92.8 102 327

induction induction induction induction (100, 3) (2.8, 4) (9.9, 4) induction (29, 3) (14, 3) (35, 3)

Di-ortho-PCBs 0.0 0 0.0 0 0.0 0 0.0 0 0.00005 5 0.001 10 0.001 10 0.0 0 0.0002 10 0.0002 10 0.00005 5

NV no induction NV NV NV NV NV no induction NV NV NV

NV NV NV NV NV NV NV NV NV NV NV

NV NV NV NV NV NV NV NV NV NV NV

NV NV NV NV NV 0.00002 0.00002 NV NV 0.00002 0.00002

NV NV NV NV NV NV NV NV NV 0.00001 NV

2,2′,3,5′,6-PeCB 2,2′,3,4,4′,6-HxCB 2,2′,3,4′,5,5′,6-HpCB

PCB 95* no induction PCB 139 29 (11, 3) PCB 187* no induction

Tri-ortho-PCBs 0.0 0 0.0006 10 0.0 0

NV NV NV

NV NV NV

NV NV NV

NV NV NV

NV NV NV

Tetra-ortho-PCBs 0.0 0 NV 0.0 0 NV 0.0 0 NV

NV NV NV

NV NV NV

NV NV NV

NV NV NV

2,2′,6,6′-TeCB PCB 54 2,2′,3,3′,6,6′-HxCB PCB 136* 2,2′,3,3′,4,4′,5,5′,6,6′-DeCB PCB 209

no induction no induction no induction

a For comparison, IEFs in rat hepatoma H4IIE cells obtained by Safe and colleagues (55, 62, 72), Tillitt and colleagues (31, 73), and Clemons and colleagues (56, 65) are shown. Also shown are International/Safe TEFs (12, 20, 21) and WHO/IPCS TEFs (14, 20, 21). PCB congeners found in herring gull and great blue heron eggs using either GC-MS or GC-ECD are indicated with an asterisk. NV, no value.

were the non-ortho-substituted congeners PCBs 126, 81, 77, and 169, with IEFs of 0.3, 0.2, 0.03, and 0.02, respectively. These congeners have chlorine atoms in both para, at least two meta, and no ortho positions, and their relatively high potencies are generally attributed to the fact that they bind with high affinity to the Ah receptor (2). As in other in vitro and in vivo systems, small changes in molecular structure resulted in remarkable differences in potency. For example, the non-ortho-substituted PCBs 78 and 81 differ only in the position of one chlorine atom, yet IEFs were approximately 2000-fold different. Among the 14 non-orthosubstituted PCBs that we tested, 12 induced EROD, with IEFs ranging from 0.000 01 to 0.3. All eight mono-ortho-

substituted PCBs that we tested induced EROD, with IEFs ranging from 0.000 02 to 0.001. Among the 11 di-orthosubstituted PCBs tested, 6 induced EROD, with IEFs of 0.000 05, 0.0002, or 0.001. PCBs 138 and 128 were as potent inducers as the monoortho-PCB 118. Of the six tri- and tetra-ortho-substituted congeners tested, only the trisubstituted PCB 139 induced EROD, with an IEF of 0.0006. PCB congeners that induced EROD caused maximal catalytic activities that ranged from approximately 5% to 100% of the maximal activity elicited by TCDD (Figure 1 and Table 1). Most of the PCB congeners with low (e0.001) IEFs had relatively low (10-30%) maximal activity, but there were exceptions. For example, PCB 78 had an IEF of

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FIGURE 1. Dose-dependent effect of 2,3,7,8-TCDD, 2,3,7,8-TCDF, and 11 PCB congeners on EROD activity in CEH cultures. Each dose-response curve was obtained on a separate 48-well plate. Points represent mean responses of triplicate doses; bars represent standard errors.

0.0001, yet its maximal activity was 90% of the maximal activity elicited by TCDD. The reason(s) for variable maximal EROD activity is not known, and several hypotheses are currently being tested in our laboratory. The phenomenon likely is not due to the insolubility of high concentrations of congeners in cell culture medium because there is a dose-dependent increase in porphyrins at the doses where EROD declines (35, 38). Variable maximal EROD activities are also observed in other cell cultures, including H4IIE rat hepatoma cells (55), fish hepatoma cells (56, 57), and primary hepatocyte cultures from several species of birds (28). A comparison of the dose-response curves of immunodetectable CYP1A and EROD in the PLHC-1 fish hepatoma cell line (57) has shown that measurement of EROD activity alone may overestimate the potency of compounds that elicit low maximal activities, but we have not observed the same phenomenon in CEH cultures (manuscript in preparation). It is important to point out the influence that variable maximal EROD activities have on the IEFs reported in Table 1. For example, PCBs 105 and 118 had IEFs of 0.005 and 0.001, yet if these compounds had elicited maximal responses similar to the response elicited by TCDD, the IEFs would have been substantially lower (assuming the same slope of the dose-response curve). Perhaps future measures of relative potency will include the concept of efficacy (58, 59) along with EC50 comparisons, but in the present paper IEFs are defined as the EC50 of TCDD divided by the EC50 of the compound to be consistent with the approach currently used by most other investigators

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working with cell culture bioassays that measure ERODbased TCDD-EQs (for example refs 29, 31, 32, 56, 60, and 61). Comparisons between CEH-Derived IEFs and H4IIEDerived IEFs. The most commonly used bioassay for measuring TCDD-EQs in environmental samples measures EROD induction in rat hepatoma H4IIE cells (29, 31, 34, 61). H4IIE- and CEH-derived IEFs are compared in Table 1. H4IIE-derived IEFs from three laboratories are shown to draw attention to some of the similarities and differences that have been reported in these cells. The CEH-derived IEF for 1,2,3,7,8-PeCDD of 1.1 was identical to the H4IIEderived IEFs reported by Clemons et al. (56), but it was 100-fold higher than the IEF reported by Safe and colleagues (62). Other differences between and within bioassays are apparent. For example, 2,3,7,8-TCDF was approximately equipotent to 2,3,7,8-TCDD in CEH cells, but has been reported to have an IEF of 0.006, 0.03, and 0.09 in H4IIE cells by Tillitt et al., Clemons et al., and Safe et al., respectively. The CEH-derived IEF for PCB 126 was identical to the IEF reported by Safe and colleagues in H4IIE cells, but was approximately 15-fold higher than the IEF reported by Tillitt et al. in H4IIE cells. There were several differences in the protocols used by the investigators working with H4IIE cells (the solvents used to dissolve congeners, cell density, etc.), and further research is required to determine the reason(s) for differences in H4IIE-derived IEFs. Regardless of interlaboratory differences in H4IIE-derived IEFs, it is important to note that CEH cells were considerably more sensitive to the non-ortho-substituted PCBs 77, 81,

FIGURE 2. Dose-dependent effect of extracts from three herring gull eggs (a) and three great blue heron eggs (b) on EROD activity in CEH cultures. HAHs were extracted from 1-g aliquots of herring gull eggs and concentrated in 300 µL of DMSO. HAHs were extracted from 5-g aliquots of great blue heron eggs and concentrated in 100 µL of DMSO. Each dose-response curve was obtained on a separate 48-well plate. Points represent mean responses of triplicate doses; bars represent standard errors. Maximal EROD activity for TCDD was 325 ( 32 pmol/min/mg of protein (n ) 3) when these extracts were analyzed.

and 169 and the mono-ortho-substituted PCBs 105, 118, 156, and 157 than were H4IIE cells. In addition, several di-ortho-substituted PCBs induced EROD in CEH cells; these congeners have not, to our knowledge, been tested in H4IIE cells. Comparisons between CEH-Derived IEFs and TEFs. CEH IEFs are compared in Table 1 to TEFs that have been proposed for human risk assessment purposes. The CEH IEF for 1,2,3,7,8-PeCDD of 1.1 was approximately 2-fold greater than its TEF, but the CEH-derived IEF for 2,3,7,8TCDF was 10-fold greater than its TEF. The TEF for TCDF takes into consideration the fact that this compound is rapidly metabolized in vivo, and the 24-h exposure period in CEH cultures may not result in significant metabolism of this compound. The non-ortho-substituted PCB congeners 77, 126, and 169 and the mono-ortho-substituted congeners 105, 118, 156, and 157 had CEH IEFs that were similar to the International/Safe TEFs. In contrast, CEH IEFs for several mono- and di-ortho-substituted PCB congeners were markedly higher than WHO/IPCS TEFs. Various biochemical and toxic end points have been used to estimate TEFs (2, 14), and they were derived from in vivo and in vitro experiments with mammalian systems. Therefore, it is not surprising to see differences between CEH IEFs and TEFs. TCDD-EQ Concentrations in Herring Gull and Great Blue Heron Eggs. Representative curves showing dosedependent effects of extracts from herring gull and great blue heron eggs on EROD activity in CEH cultures are shown in Figure 2. As found with all EROD-inducing PCB congeners, the curves were bell-shaped, and the maximal activity elicited by extracts was lower than the maximal activity elicited by TCDD: 325 ( 32 pmol/min/mg of protein. We considered the possibility that biogenic compounds coextracted from egg homogenates along with PCB, PCDD/Fs, and other HAHs might diminish activity, but maximal EROD activities and EC50’s of complex mixtures of PCBs (Aroclors 1248 and 1254) or individual PCBs were identical to curves obtained from extracts prepared from PCB-spiked chicken eggs. Therefore, coextracted biogenic compounds appeared to have little or no influence on maximal EROD activity.

Equation 2 was used to obtain TCDD-EQbio concentrations, and eq 3 was used to obtain TCDD-EQcalc concentrations from chemical residue data in conjunction with CEH IEFs. A linear regression of TCDD-EQbio against CEHderived TCDD-EQcalc concentrations is shown in Figure 3A (top). The regression line (slope, 0.99; r2 ) 0.977; p < 0.0001) was forced through 0 because extracts from uncontaminated eggs (commercial chicken eggs) showed no TCDDEQbio. The regression line not forced through 0 had a y-intercept of 2.9 and the slope was 0.96. TCDD-EQ concentrations were lower in great blue heron eggs than in herring gull eggs. Herring gull eggs that had TCDDEQbio concentrations equal to or greater than 25 ng/g were obtained from the early to mid-1970s, a period of time when the Great Lakes were considerably more contaminated with PCBs, PCDD/Fs, and other organochlorines than they are currently (45, 63). The center and lower panels of Figure 3A show the percent contributions of PCB congeners, 2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, and 2,3,7,8-TCDF to TCDDEQcalc when using CEH-derived IEFs. PCBs contributed approximately 99.5% to CEH-derived TCDD-EQcalc concentrations in herring gull eggs, with the major contributors being PCBs 138, 105, 126, 66, and 118. Although 2,3,7,8TCDD, 1,2,3,7,8-PeCDD, and 2,3,7,8-TCDF were minor contributors (