Aryl hydrocarbon receptor-mediated activity of gas-phase ambient air derived from passive sampling and an in vitro bioassay

Passive air sampler deployment

Low-density polyethylene sheeting (United Plastics), 800-μm-thick, was cut into approximately 7.5- × 13-cm pieces and cleaned in solvent (dichloromethane and hexane) to remove background contamination. At each of 9 sampling sites throughout the Cleveland area, 4 polyethylene passive samplers were fastened inside an inverted stainless steel bowl using zip-ties and the bowl was suspended so that the samplers were hanging at a height of approximately 2 m.

To calculate ambient air concentrations from concentrations measured in deployed polyethylene passive samplers, performance reference compounds (PRCs) are often added to the samplers for in situ calibration of sampling rates. However, PRCs could not be added to the samplers intended for bioassays because these compounds would interfere with bioassay response. Therefore, 50-µm-thick polyethylene passive samplers, preloaded with PRCs by incubation in an 80:20 methanol:water solution, were codeployed at each site, and sampling rates determined for these 50-µm samplers were used to interpret results from 800-µm samplers.

A map of the study region is shown in the Supplemental Data (Figure S1), and characteristics of the deployment sites are summarized in Table 1. Deployments took place from June to September of 2013, with each set of samplers deployed for approximately 60 d. After deployment, the samplers were removed from the protective bowl, wrapped in precombusted aluminum foil, and shipped on ice overnight to the University of Rhode Island Graduate School of Oceanography (Narragansett, RI, USA), where they were kept frozen until extraction.

Sample preparation

Each 800-µm polyethylene passive sampler was extracted twice in pentane, each time for 18 to 24 h. The 50-µm samplers were extracted once for 18 to 24 h in pentane. Every batch of samplers was extracted along with a laboratory blank, which was a polyethylene passive sampler that had been cleaned alongside the field samples and then stored frozen in precombusted aluminum foil. All 4 800-µm samplers deployed simultaneously at the same site were composited into one extract and concentrated to 1 mL in a warm water bath under a gentle stream of nitrogen. Extracts from the 800-µm samplers appeared to contain a white precipitate, possibly from co-extracted polyethylene material. To remove the particulate, extracts were serially frozen, causing the precipitate to solidify at the bottom of the vial, and the overlying liquid was removed via a Pasteur pipette and reconstituted to 1 mL with pentane. Two aliquots were removed from the 1-mL solution, one for chemical analysis and the other for biological analysis. A schematic summarizing sample preparation is shown in the Supplemental Data (Figure S2).

Chemical analysis by GC–MS

The fraction of polyethylene passive sampler extract intended for chemical analysis was spiked with the internal standards acenaphthene-d₁₀, phenanthrene-d₁₀, chrysene-d₁₂, and perylene-d₁₂ and analyzed on an Agilent 6890 GC device coupled to an Agilent 5973 Mass Selective Detector (MSD) in electron impact (70 eV) mode for 22 PAHs, 18 alkylated PAHs, and (in a separate GC–MS run) 12 OPEs using an Agilent J&W DB-5 fused capillary column (30 m × 0.25 mm I.D.). The PAHs were quantified using an 8-point calibration curve with linearity r² > 0.990 for all compounds. The OPEs were quantified using a 10-point calibration curve with linearity r² > 0.997 for all compounds except TDBPP, which was not detected in samples and is omitted from further discussion.

Extracts were also spiked with non-native polybrominated diphenyl ethers (BDEs 35, 77, 128, and 183) and analyzed on an Agilent 7890 GC device coupled to an Agilent 5977 MSD in negative chemical ionization mode with methane reagent gas for 12 BDEs and 8 novel halogenated flame retardants (NHFRs), as well as 3 polybrominated biphenyls, which were used as PRCs in sampling rate determination for codeployed thin polyethylene passive samplers. A complete list of target compounds and abbreviations is available in the Supplemental Data (Table S1). The BDEs and NHFRs were quantified using an 8-point calibration curve with linearity r² > 0.995.

To avoid interference with biological assays, samples were not spiked with internal standard prior to extraction and so were not corrected for internal standard recoveries. Concentrations presented for polyethylene passive sampler extracts were not blank-subtracted before use in data interpretation. This was considered appropriate because our primary interest was in determining the actual concentration present in the bioassay exposure solution.

Calculation of ambient air concentrations

The composition of HOCs accumulated in polyethylene differs from the ambient composition of gas-phase HOCs in air because the concentration in polyethylene is dependent not only on gas-phase concentrations, but also on the affinity of each compound for the polyethylene passive sampler matrix and the rate at which the compound is absorbed into the sampler. To compare the composition of solutions used in bioassay experiments with the actual composition of gaseous HOCs expected in ambient air, gaseous HOC concentrations were calculated based on the results of chemical analysis of polyethylene passive sampler extracts and sampling rates determined from codeployed PRC-loaded polyethylene passive samplers. Concentrations were blank-subtracted using the co-extracted laboratory sampler blank. After blank subtraction, concentrations <25% of the sampler blank were considered <detection limit, and all <detection limit values were replaced with 0.

To translate concentrations within the polyethylene passive sampler to concentrations in ambient air, the volume of air sampled by each sampler during deployment was estimated using data on the percentage of loss of labeled PRCs from codeployed 50-µm-thick samplers. From the PRC loss data, the best-fit value for the thickness of the diffusive boundary layer (DBL) at the air–polyethylene passive sampler interface was determined. Because all the samplers were deployed under the same conditions and the thickness of the sheet does not affect air-side resistance, the DBL thickness determined for thin sheets was then used in a 2-film model describing sampler-side and air-side mass transfer rates to calculate the percentage of equilibration reached by each target compound in the 800-µm-thick samplers. This approach for estimation of percentage of equilibration from PRC loss data has been described in detail in previous work (McDonough et al. 2016).

Biological analysis by reporter cell bioassay

Aliquots for biological analysis were mixed with 200 µL of dimethyl sulfoxide (DMSO) and blown down under a gentle stream of nitrogen to constant volume. This stock solution was then used to create a 10-point dilution curve (0.01–120 g polyethylene passive sampler/mL) for each sample, including the sampler blank (Supplemental Data, Figure S1).

The AhR reporter cell line used was H1G1.1c3, a murine hepatoma cell line consisting of Hepa-1c1c7 cells stably transfected with AhR-responsive green fluorescent protein (GFP) reporter gene (Nagy et al. 2002). Cells were plated in 96-well plates (3 × 10⁵ cells/well; Costar 96-well black plate with a clear bottom) and allowed to attach overnight at 37 °C in selective medium (Nagy et al. 2002). The medium was then changed to nonselective medium, and the cells in 100 μL of medium were treated with 1 µL of each sample dilution for a final vehicle concentration of 1% DMSO. All wells were prepared in triplicate and incubated at 33 °C. For each test extract, the cells in 3 wells were treated with 1 µL of DMSO as a negative control, and the cells in another set of 3 wells were left untreated to control for any natural cell fluorescence. On each plate, 3 wells were treated with BaP at a final well concentration of 120 nM dissolved in DMSO as a positive control. On one plate, a 10-point dilution curve was also run for BaP (1.2 × 10⁻⁵–12 000 nM), and the results were normalized to the positive control 120 nM BaP (Supplemental Data, Figure S3).

Activity mediated by AhR was measured by reading the GFP fluorescence emitted by the cells at 515 nm using a Spectra Max M3 plate reader at 24 and 48 h post dosing (hpd). The mean fluorescence value of the DMSO-treated negative control triplicate wells was subtracted from each sample's fluorescence reading, and the response was expressed as a ratio over the mean fluorescence value for the triplicate 120-nM BaP-positive controls run on the same plate to control for plate-to-plate differences in cell response.

Calculation of extract potency

Data from 48-hpd readings were fitted to a 4-parameter log–logistic concentration–response model with the lower bound set to 0 using the R package drc (Ritz et al. 2015). The response f occurring as a result of concentration x is modeled as in Equation 1, where c is the lower bound value (set to 0), d is the upper bound value, b determines slope steepness, and e is the concentration achieving 50% of maximum efficacy (EC50). The upper bound was set to the maximum observed response in cases in which the response reached a plateau or decreased at the highest dosages, but was not defined for the extract from site Cleveland Lakefront 1 because the response continued increasing up to the maximum extract concentration.

(1)

In addition to the EC50, the EC_BaP50 was calculated as an alternative measure of potency. The EC_BaP50 is the concentration resulting in 50% of the effect observed for the plate-specific positive control (120 nM BaP). The EC_BaP50 was identified as a more useful metric than EC50 because the concentration–response curves of the extracts were not parallel and maximum efficacy varied among curves.

Dosing solutions were prepared so that each sample was representative of the same amount of extracted polyethylene passive sampler to facilitate comparison with the sampler blank and control for any interference caused by background contamination in the sampler matrix. However, due to site-to-site variability in sampling rates, the volume of air represented by each sample differed among sites (Table 1). For this reason, after determination of EC_BaP50 from the concentration–response curve fit, EC_BaP50 values were normalized based on the volume of air sampled at each site. Aliquots of polyethylene passive sampler extracts used in dosing solutions were representative of 1900 to 3100 m³ of air, and were all normalized to 2000 m³.

To compare predicted AhR-mediated potency based on chemical composition with observed potency based on bioassay experiments, BaP equivalents were calculated for both sets of data. For concentrations measured via chemical analysis, BaP equivalents in each mixture (BaP-Eq_chem) were determined as in Equation 2 by multiplying the concentration of each compound in the polyethylene passive sampler extract (C_n) by the compound's potency relative to BaP (expressed as IEF_n) using values from Machala et al. (2001) and summing results for all compounds. Benzo[b]fluoranthene, benzo[k]fluoranthene, and benzo[j]fluoranthene could not be quantitated separately with confidence via the chromatographic method used, so the IEFs for these 3 compounds were averaged as an estimate of the IEF for benzo[b,j,k]fluoranthene. Concentrations in the polyethylene passive sampler extract were normalized based on volume of air sampled before BaP-Eq_chem calculations were done.

(2)

For the bioassay results, the BaP equivalent of each sample extract was expressed as the amount of BaP needed to achieve the same response as the extract. The bioassay-derived toxic equivalency (BaP-Eq_bio) was calculated as in Equation 3 as the ratio of the amount of BaP needed to achieve a response of 50% of the maximum efficacy (EC50 for the BaP curve) over the volume of polyethylene passive sampler extract added to the well to achieve that same effect (the EC_BaP50 of the extract).

(3)

The degree to which chemical analysis explained observed potency (%_chem) was then expressed as in Equation 4.

(4)

Chemical composition of passive sampler extracts

Concentrations of all compounds in polyethylene passive sampler extracts are presented in the Supplemental Data (Table S2) for PAHs, OPEs (Supplemental Data, Table S3), and halogenated flame retardants (HFRs; Supplemental Data, Table S4). Concentrations of PAHs and OPEs in the sampler extracts are displayed in Figure 1A and B along with estimated ambient air concentrations (Figure 1C and D). All concentrations for field samples were normalized to an air volume of 2000 m³ to facilitate comparison between sites.

Total alkyl and parent PAHs (Σ₄₀PAH) in polyethylene passive sampler extracts ranged from 3.6 ng/µL for the extract from Cuyahoga National Park to 34 ng/µL for a residential suburban area in University Heights. Concentrations of PAHs were dominated by phenanthrene (0.6–16.3 ng/µL; 10–57%), fluoranthene (0.1–6 ng/µL; 1–18%), 2-methylphenanthrene (0.1–1 ng/µL; 1–6%), and fluorene (0.3–1 ng/µL; 3–9%).

Concentrations of OPEs were much greater than those of HFRs. Total OPEs (Σ₁₂OPE) ranged from 0.4 ng/µL for the extract from Cuyahoga National Park to 2.0 ng/µL for a residential area in Kent. The Σ₁₂OPE was dominated by TPHP at all downtown Cleveland sites (0.09–0.78 ng/µL; 28–69%), whereas Cuyahoga National Park and Fairport Harbor were dominated by TEHP (0.30–0.57 ng/µL; 68%), and University Heights and Kent were dominated by tri-n-butyl phosphate (TNBP; 0.50 ng/µL; 59%) and TCIPP (1.60 ng/µL; 78%), respectively. Concentrations of total BDEs (Σ₁₂BDE) ranged from 10 pg/µL in Cuyahoga National Park to 46 pg/µL at Downtown Cleveland Site 2, and were dominated by BDE 47 and 154. Concentrations of total NHFRs (Σ₁₈NHFRs) were greatest in the polyethylene passive sampler blank due to the presence of 1,2-bis(2,4,6-tribromophenoxy)ethane and Dechlorane Plus, which were not found in any of the field sample extracts.

The 2-, 3-, and 4-ring PAHs, as well as their alkylated and substituted counterparts, were generally correlated in the different extracts (0.3 < r² < 0.9), whereas the 5- to 6-ring PAHs did not exhibit significant correlation among themselves or with any other group of PAHs (Supplemental Data, Table S5). Correlation among individual PAHs was expected, because they are typically emitted from the same sources. Correlation among PAHs was further confirmed by principal component analysis, which showed that 76% of variation in samples was explained by 2 principal components, the first with loadings primarily from 3- to 4-ring PAHs, and the second with loadings primarily from 2- and 4- to 5-ring PAHs (Supplemental Data, Figure S4). In contrast, individual OPEs were generally not significantly correlated, although some degree of correlation (r² ≥ 0.3) was observed between TDCIPP and TNBP (Supplemental Data, Table S6). In addition, TNBP, TDCIPP, and ethylhexyl diphenyl phosphate (EHDPP) exhibited some correlation with PAHs (Supplemental Data, Table S7).

Ambient air concentrations

Ambient gaseous concentrations of Σ₄₀PAH ranged from 7.1 ng/m³ in Cuyahoga National Park to 36.2 ng/m³ at urban site Cleveland Downtown 1 and were dominated by the methylnaphthalenes (1.7–8.8 ng/m³; 18–33%), phenanthrene (0.3–9.8 ng/m³; 2–33%), and fluorene (0.5–2.6 ng/m³; 5–14%). Concentrations were similar in range to those measured by Peverly et al. (2015) in Chicago using polyurethane foam passive samplers in 2012 to 2014 (Σ₁₆PAH = 9–52 ng/m³), and by Melymuk et al. (2012) in Toronto in 2007 to 2008 (Σ₂₇PAH = 0.3–51 ng/m³), also using polyurethane foam passive samplers. Concentrations in the present study were similar but lower than previous measurements of total gaseous PAHs using polyethylene passive samplers in the downtown Cleveland area by McDonough et al. (2014) in 2012 (Σ₁₅PAH = 23–80 ng/m³). In larger scale regional studies, atmospheric concentrations of PAHs have often been found to correlate with population density (Hafner et al. 2005; McDonough et al. 2014), but in the present study no significant (p < 0.05) correlation between gaseous PAH concentrations and population density within 5 to 30 km was observed.

Gaseous concentrations of Σ₁₂OPE ranged from 0.01 ng/m³ in Cuyahoga National Park to 1.1 ng/m³ in Kent. This was similar in range to measurements by Peverly et al. (2015) in Chicago using polyurethane foam passive samplers in 2012 to 2014 (Σ₁₃OPE = 0.5–1.5 ng/m³), and slightly lower than measurements of particulate Σ₁₂OPE in the Cleveland area by Salamova et al. (2014) in 2012 (mean Σ₁₂OPE = 2.1 ± 0.4 ng/m³. The most abundant OPE at all sites was TCIPP (0.01–1.0 ng/m³; 9–98%) except at University Heights, where TNBP dominated (0.6 ng/m³; 87%). In a previous study, TCIPP was also found to be the most abundant in Cleveland particulate Σ₁₂OPE (0.85 ± 0.3 ng/m³; Salamova et al. 2014).

Figure 1 compares the chemical composition of the polyethylene passive sampler extracts used in bioassay experiments with that of ambient gaseous PAHs and OPEs. Extracts used in bioassays were enriched in moderately hydrophobic compounds, such as fluoranthene and TDCIPP, which make up a lower percentage of total HOCs in the gaseous fraction of ambient air but have a greater affinity for the polyethylene passive sampler matrix. The different HOC compositions in the gas phase and in the sampler extract illustrate that it is not possible to estimate the total AhR-mediated potency of the mixture that is present in gas-phase air. However, AhR-mediated potency results based on the HOC mixture found in polyethylene passive sampler extracts still represent an important step in understanding the biological relevancy of gas-phase compounds. Furthermore, the composition in polyethylene passive sampler extracts is expected to be more similar to the composition of HOCs diffusing into plant material or skin from air, or accumulating in the body via other mechanisms.

Concentration–response curves

Extracts from all polyethylene passive samplers, including the sampler blank, induced concentration-dependent activation of AhR-dependent GFP. All concentration–response data are displayed along with curve fits and 95% confidence intervals in Figure 2, with response represented as a ratio compared with response elicited by the plate-specific positive control. For all extracts, an initial increase in GFP induction was seen with increasing concentration. However, there was a precipitous decline in the fluorescence for all extracts (except Cleveland Lakefront 1) at the greatest concentrations, possibly due to cytotoxicity or inhibition of fluorescence response at high concentrations of polyethylene passive sampler extract. These points were omitted during concentration–response curve fitting, because we were interested in determining only the induction potencies of the extracts. Most extracts did not exhibit a clear plateau in response, making determination of maximum efficacy, as well as EC50, somewhat uncertain. Furthermore, maximum efficacy of the samples varied from 94 to 230% of positive control response (Table 2). For this reason, EC_BaP50, measured relative to the plate-specific positive control, was used to compare the potencies of the samples.

The EC_BaP50 of each extract, normalized based on the volume of air sampled at each site, is displayed in Table 2 along with each extract's maximum observed efficacy. Values of EC_BaP50 ranged from 0.5 ± 0.1 g polyethylene passive sampler/mL at Downtown Cleveland 1 to 6.6 ± 1.2 g polyethylene passive sampler/mL at Cuyahoga National Park.

The 3 rural/residential sites had the lowest potency (greatest EC_BaP50 values), ranging from 2.6 to 6.6 g polyethylene passive sampler/mL, followed by the 2 Cleveland Lakefront sites. The most potent extracts were from the 3 Cleveland Downtown sites and one semi-urban residential site (University Heights, a densely populated suburb). This contrasts with work by Klein et al. (2005), who found no change in potency of gaseous extracts between urban and rural samples with distinct chemical compositions, but is consistent with work by Érseková et al. (2014), who found that extracts from impacted sites were more potent in AhR bioassays than extracts from rural sites. The potency of the polyethylene passive sampler blank (EC_BaP50 = 23 ± 5 g polyethylene passive sampler/mL) was significantly lower than that of all field samples. Blank comparisons were done before normalizing for the volume of air sampled so that each sample would be representative of the same mass of extracted polyethylene.

The potency and maximum efficacy of the extracts did not appear to be correlated. This is most likely due to a complex interplay between the unique composition of ligands in each sample, their affinity for the AhR, the resulting ligand–receptor complex's ability to bind other necessary transcription factors, and cytotoxicity of specific components. Response could also be affected by ligands interacting with other pathways that could amplify or dampen AhR response. Klein et al. (2005) also observed a lack of correlation between potency of extracts and maximum efficacy with respect to AhR binding of gas-phase extracts from active air sampling.

Initial bioassay experiments demonstrated that the fluorescence responses of the treated cells increased over time from 16 to 48 hpd, so all responses reported here were measured at 48 hpd. This is in contrast to other studies of AhR activation for environmental samples, most of which have used a luciferase reporter rather than the GFP reporter we used. For example, Machala et al. (2001) measured the greatest potency at 6 hpd, most likely due to PAH metabolism, and Kennedy et al. (2010) observed steadily decreasing potency in extracts from 24 to 72 hpd. This discrepancy is most likely due to differences in induction kinetics and increased stability of the GFP reporter compared with the luciferase reporter (Han et al. 2004). It is also possible that some of the response observed in this study was due to compounds that were less readily metabolized than PAHs and OPEs.

Bioassay-derived BaP equivalents for polyethylene passive sampler extracts

Figure 3 shows a map of results for BaP-Eq_bio alongside maps of total concentrations of PAHs and OPEs in the polyethylene passive sampler extracts (Σ₄₀PAH and Σ₁₂OPE). The BaP-Eq_bio values ranged from 21 to 283 ng/µL BaP equivalents and were generally greatest in the downtown Cleveland area and lowest at the rural/residential sites further from the city center.

The BaP-Eq_bio values were compared with concentrations of PAHs and organic flame retardants (OPEs, PBDEs, and NHFRs) in the polyethylene passive sampler extracts to determine whether there was any significant correlation between potency and chemical composition. Although some correlations were found, few were likely to be driving potency. No correlations with PBDE and NHFR concentrations were observed. The BaP-Eq_bio values weakly correlated only with 2-ring alkyl/substituted PAHs (r² = 0.42; p < 0.1; standard error [SE] = 64; n = 9) and also displayed some correlation with EHDPP (r² = 0.66; p < 0.01; SE = 49; n = 9). Maximum efficacy of polyethylene passive sampler extracts showed some correlation with concentrations of 3-ring (r² = 0.61; p < 0.05; SE = 31; n = 9) and 4-ring (r² = 0.48; p < 0.05; SE = 36; n = 9) parent PAHs. Correlations between BaP-Eq_bio values and alkyl/substituted PAHs were only investigated by grouping compounds (2-ring alkyl/substituted PAHs; 3–4-ring alkyl/substituted PAHs) because quantitative standards were not available for all alkylated PAHs. However, it is important to note that AhR-mediated potency differs greatly between PAH isomers. Because a high degree of correlation was observed between different low-molecular-weight PAHs at different locations in the present study (Supplemental Data, Table S5), it was expected that the composition of alkyl/substituted PAHs would most likely be similar between sites, so correlations with BaP-Eq_bio values are likely driven by the same compounds at all sites.

Little information is available regarding the biological effects of alkylated PAHs. Recent studies using a yeast reporter assay system and an H4IIE-luc reporter-gene assay suggest that methyl- and dimethyl-substituted phenanthrenes are in some cases more potent with respect to AhR activation than their unsubstituted counterparts (Sun et al. 2014; Lam et al. 2018). The statistically significant correlation between BaP-Eq_bio and EHDPP suggests that this compound, or unmonitored compounds with which it covaries spatially, could be contributing to AhR activity. Because no compelling evidence is available for EHDPP as an AhR activator, the presence of other AhR activators that covary with EHDPP is somewhat more likely. Previous studies have shown that levels of OPEs and other OFRs can correlate in air due to their historical use in the same formulations (Salamova et al. 2014). In addition, some OPEs that were not targeted in this study, including mono-substituted isopropyl triaryl phosphate (mITP), have been shown to have relatively strong AhR activity (Gerlach et al. 2014; Haggard et al. 2017).

Predicted BaP equivalents from chemical analysis

The BaP-Eq_chem of each polyethylene passive sampler extract was calculated based on concentrations of targeted PAHs from GC–MS analysis. No dataset for the specific cell line we used was available, so IEFs were taken from Machala et al. (2001), who measured PAH-induced AhR-mediated response in a rat hepatoma H4IIE cell line stably transfected with luciferase reporter gene. The IEFs were not available for all PAHs, so calculated BaP-Eq_chem values are representative of only 14 compounds (Supplemental Data, Table S8). Although the dataset from Machala et al. (2001) is the most applicable that could be found, these IEFs come from a cell line with a completely different time-dependent expression profile and are not directly applicable to the cell line we used. This contributes greatly to the uncertainty in the derived BaP-Eq_chem values, and highlights the need for more studies providing cell–line-specific IEFs for a wide range of ubiquitous environmental contaminants.

The BaP-Eq_chem values calculated using potencies from Machala et al. ranged from 1.6 to 7.9 ng/µL BaP, as shown in Table 3. The percentage of BaP-Eq_bio accounted for by this BaP-Eq_chem is also displayed. The percentages of contributions of individual PAHs to the total predicted BaP-Eq_chem are displayed in Figure 4. Among the targeted PAHs, contributions to BaP-Eq_chem were dominated by high-molecular-weight PAHs that were present at low concentrations in the polyethylene passive sampler extracts, including dibenz(a,h)anthracene, indeno(1,2,3-c,d)pyrene, benzo(b/k)fluoranthene, and chrysene.

Potencies calculated from known chemical composition using IEFs explained only 2 to 23% of the AhR-mediated potency observed in bioassay experiments (Table 3), and BaP-Eq_chem and BaP-Eq_bio were not significantly correlated. This finding suggests that other compound groups present in the gaseous fraction of ambient air may also be contributing to BaP-Eq_bio of the extracts. These may include additional parent PAHs and alkyl-PAHs not measured in the present study, as well as oxygenated PAHs and N- and S-heterocyclic PAHs (Larsson et al. 2014; Sun et al. 2014; Lam et al. 2018). Compounds other than PAHs may also be responsible for some of the observed AhR-mediated potency. The use of BaP-Eq_chem values derived from a different bioassay may also contribute to this discrepancy.

The correlation observed between concentrations of EHDPP and AhR activity suggests that this compound, or other OFRs with a similar source, may be contributing to BaP-Eq_bio as well, although further research is needed to understand the AhR-mediated potency of OFRs. Furthermore, a major weakness of predicting potency based on compound IEFs is that it considers only additive interactions, without taking into account synergistic and antagonistic effects, which are highly probable in complex environmental mixtures. This factor, along with the scarcity of IEF values for the targeted compounds, most likely contributed to the discrepancy between observed and predicted AhR-mediated potency.