Do 16 Polycyclic Aromatic Hydrocarbons Represent PAH Air Toxicity?

Estimation of carcinogenic potency based on analysis of 16 polycyclic aromatic hydrocarbons (PAHs) ranked by U.S. Environmental Protection Agency (EPA) is the most popular approach within scientific and environmental air quality management communities. The majority of PAH monitoring projects have been focused on particle-bound PAHs, ignoring the contribution of gas-phase PAHs to the toxicity of PAH mixtures in air samples. In this study, we analyzed the results of 13 projects in which 88 PAHs in both gas and particle phases were collected from different sources (biomass burning, mining operation, and vehicle emissions), as well as in urban air. The aim was to investigate whether 16 particle-bound U.S. EPA priority PAHs adequately represented health risks of inhalation exposure to atmospheric PAH mixtures. PAH concentrations were converted to benzo(a)pyrene-equivalent (BaPeq) toxicity using the toxic equivalency factor (TEF) approach. TEFs of PAH compounds for which such data is not available were estimated using TEFs of close isomers. Total BaPeq toxicities (∑88BaPeq) of gas- and particle-phase PAHs were compared with BaPeq toxicities calculated for the 16 particle-phase EPA PAH (∑16EPABaPeq). The results showed that 16 EPA particle-bound PAHs underrepresented the carcinogenic potency on average by 85.6% relative to the total (gas and particle) BaPeq toxicity of 88 PAHs. Gas-phase PAHs, like methylnaphthalenes, may contribute up to 30% of ∑88BaPeq. Accounting for other individual non-EPA PAHs (i.e., benzo(e)pyrene) and gas-phase PAHs (i.e., naphthalene, 1- and 2-methylnaphthalene) will make the risk assessment of PAH-containing air samples significantly more accurate.

Many studies [35] have used the 16 U.S. EPA PAHs due to their recognized toxicity by the scientific community in the late 1960s and early 70s. Analytical capabilities and standard materials available at that time imposed some limitations on the selection of PAHs and thus further health risk assessment. At present, a larger number of PAH compounds can be analyzed simultaneously. That opens a discussion for whether other PAH compounds should be recognized as priority PAHs and studied for their toxic properties.
This paper presents an analysis of results from a unique set of 13 projects in which 88 PAHs were measured in both particle and gas phases. PAH samples were collected from different types of sources (biomass burning, vehicle emissions, meat cooking, mining operations and urban air). The goal of this study was to compare TEF-based carcinogenic potency of the 16 U.S. EPA PAHs with the total TEF-based carcinogenic potency of the measured 88 PAHs in both phases and to determine whether measurements of only the 16 U.S. EPA particle-bound PAHs adequately represent the total potential toxicity of PAHs originating from different sources. Using the TEF approach, the relative importance of gaseous and particulate PAHs, as well as the most prominent contributors to the total PAH toxicity in different sources were determined.

Materials and Reagents
Certified PAH standards were purchased from Sigma-Aldrich (St. Louis, MO, USA), AccuStandard (New Haven, CT, USA), and Cambridge Isotope Laboratories, Inc. (Andover, MA, USA). Acetonitrile, acetone, methanol, and dichloromethane (High-Performance Liquid Chromatography grade) were obtained from Fisher Scientific (Fair Lawn, NJ, USA). TIGF filters (Fiber Film T60A20, PALL Life Sciences, Ann Arbor, MI, USA) and XAD-4 resin (Amberlite ® XAD4, St. Louis, MO, USA) were used as sampling media. Prior to sampling, XAD-4 resin was cleaned using an accelerated solvent extractor (ASE, DIONEX, ASE-300, Salt Lake City, UT, USA) with methanol followed by dichloromethane for 15 min each at 85 • C and 1500 psi pressure. Clean XAD-4 resin was dried in a vacuum oven at room temperature overnight. Glass (Tisch Environmental, Inc., Village of Cleves, OH, USA) and stainless-steel Teflon-coated cartridges (homemade DRI's cartridges) were packed with 20-200 g of XAD-4 resin depending on the project (Table 1) and expected concentration levels of PAHs. Every 10th blank XAD-4 cartridge and TIGF filter were extracted on ASE and analyzed with gas chromatography mass spectrometry (GC-MS) to determine their purity. Ultra-high purity (UPH) grade nitrogen was purchased from Airgas (Sparks, NV, USA).

Sampling
Gas-and particle-phase PAH concentrations were measured during several projects aimed at the characterization of: urban ambient air (A-1, A-2), biomass burning (B-1), meat cooking (M-1), engine exhaust (E1-E5), traffic emissions in a road tunnel (T-1, T-2), and mining operations (Mi-1, Mi-2) ( Table 1). In total, 325 filter and XAD samples (650 individual samples) were collected, extracted, and analyzed. Detailed descriptions of sampling campaigns can be found in previously published papers and reports (Table 1). Briefly, particulate matter (PM) was collected using medium and high-volume samplers at different flow rates depending on the project (Table 1). TIGF filters (Fiber Film T60A20, PALL Life Sciences, Ann Arbor, MI, USA) were used to collect PM for PAH analysis. Gas-phase semi-volatile PAHs were collected with XAD-4 resin cartridges that were placed downstream of the filters.
After sampling, each filter and XAD cartridge was packed into pre-cleaned aluminum foil, sealed in secure static-sensitive double-track zipper bags (Uline, Pleasant Prairie, WI, USA), and transported in a cooler with blue ice to DRI, where it was stored at −20 • C until extraction and analysis. A detailed description of the extraction and GC-MS analysis are described elsewhere [13,79].

Calculation of BaPeq Toxicity
To estimate the carcinogenic potency of analyzed PAHs, TEF values proposed by Nisbet and LaGoy [51] were applied. TEFs are widely used to calculate toxicities of PAH mixtures based on the toxicity of a specific PAH compound relative to the toxicity of the reference compound-benzo(a)pyrene. The potential cancer risk associated with exposure to PAH mixtures is calculated by summing up the concentrations of individual PAH compounds, multiplied by their TEF values (Table S1). TEF values were not available for all 88 PAHs measured in this study. TEFs have been determined for only 17 PAHs [51], including the 16 EPA priority PAHs [acenaphthylene, acenaphthene, fluorine, phenanthrene, anthracene, fluoranthene, pyrene, benz(a)anthracene, chrysene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, dibenz(a,h)anthracene, benzo(ghi)perylene, indeno(1,2,3-cd)pyrene), and 2-methylnaphthalene (Table S1)]. TEFs for the remaining 71 PAHs were estimated based on the similarity of a compound to one of the 17 PAHs whose TEFs are known. Compounds were considered similar if they are close isomers, have the same number of aromatic rings, GC-MS response factors, and similar molecular structure. For example, for 1-methylnaphthalene, the TEF of 2-methylnaphthalene (0.001) [51] was used since these compounds are isomers and are structurally similar. The same TEF (0.001) was applied for dimethyl-and trimethylnaphthalenes, because GC-MS response factors of these PAHs are similar to methylnaphthalene and because chemical structures of these compounds differ only by the presence of one or two methyl (-CH 3 ) groups. The TEF proposed for anthracene (0.01) was used to calculate the BaPeq concentrations for methylanthracenes and different benz(a)anthracenes like 7-methylbenz(a)anthracene and 7,12-dimethylbenz(a)anthracene. In the case of fluoranthenes [1 + 3-methylfluoranthene, 4-methylfluoranthene, benzo(ghi)fluoranthene, etc.], TEF of fluoranthene (0.001) was applied. TEFs of four isomers of dibenzopyrene [dibenzo(a,e)pyrene (TEF: 1), dibenzo(a,l)pyrene (TEF: 10), dibenzo(a,i)pyrene (TEF: 10), dibenzo(a,h)pyrene (TEF: 10)] were assigned using values provided by Office of Environmental Health Hazard Assessment (OEHHA) report [80]. It has to be emphasized that TEFs assigned based on analytical criteria described above may not adequately reflect the real toxicological properties of non-EPA PAHs. Since to our knowledge, there are no toxicological studies on these compounds, we used this approach to estimate the relative importance of the non-EPA PAHs and selected those that deserved a more thorough toxicological characterization. It should be also noted that 75% of the assigned TEFs (Table S2) had the lowest TEF value (0.001), which makes overestimation of their toxicity less likely. Table S2 summarizes TEFs that were assigned to the 88 PAHs measured in this study. BaPeq concentrations of PAH mixtures in different samples were then calculated using the following formula: where Ci-concentration of PAH compound i (in ng m 3 ); TEFi-assigned toxic equivalency factor (Table S2). Total BaPeq concentrations of 16 EPA priority PAHs (∑ 16 BaPeq) were calculated using the same formula, but this was applied only to the 16 EPA PAHs [51].

Results and Discussion
Measured PAH concentrations varied significantly from project to project, because of the differences in the nature of each sampling project (see Experimental Section). For example, the concentration of gas-phase naphthalene measured during controlled combustion of five different fuels [78] was 2637 ± 929 ng m −3 on average, while its concentration in ambient urban air in downtown Reno (NV, USA) was more than 10 times lower (167 ± 71 ng m −3 ). To estimate the carcinogenic potency of PAH-containing samples (Table 1), the sum of BaPeq concentrations of 88 (∑ 88 BaPeq) and 16 (∑ 16 BaPeq) gas-and particle-phase PAHs, respectively, were calculated for each project and summarized in Table S3. As expected, ∑BaPeq values vary significantly from project to project because of the differences in PAH concentrations observed during each project (Table 1). For example, the mean ∑ 88 BaPeq value for particle-phase PAHs during the biomass-burning project was 41.3 ± 22.1 ng m −3 , which was significantly higher than that of the measured urban PM (project A-1), where the ∑ 88 BaPeq was 0.341 ± 0.242 ng m −3 .
To compare the total BaPeq toxicity of the 16 EPA particle-phase PAHs with that of 88 PAHs (∑ 88 BaPeq), as well as with BaPeq toxicities that include gas-phase PAHs, the following ∑BaPeq ratios were calculated for each sample collected during the 13 projects considered here (Table S4): The above ratios helped us to assess whether the 16 EPA particle-bound PAHs were sufficient for evaluation of the total PAH toxicity of aerosols originating from different sources. Ratios close to 1 would indicate that the 16 EPA PAHs adequately represented the total toxicity, while lower ratios will provide an estimate of how much the PAH toxicity is underestimated by the traditional 16-PAH method. Summary statistics of these ratios for each project are given in Table S4. For simplicity, only the mean values of the ∑BaPeq ratios will be discussed in the text below. Figure 1 presents a graphical summary of BaPeq(%) I ratios [Equation (2)]. BaPeq(%) I that considered only particle-phase PAHs did not exceed 60% in any of the analyzed projects. The particle-bound 16 EPA priority PAHs were found to be responsible, on average, only for 21-58% of the PM PAH toxicity (Table S4). The lowest project-average BaPeq(%) I (~21%) was observed for heavy-duty vehicle engine emissions (project E-5, Table S4). It should be noted that within each project there were BaPeq(%) I observations that were significantly lower than the project-average values. These results indicated that PAHs not included in the EPA priority list contributed more to the PM toxicity than the 16 EPA PAHs, especially in the case of heavy-duty vehicle emissions (E-projects, Table 1 and S4). Analysis of BaPeq values for individual 88 PAHs showed the dominance of dibenzo(a,i)pyrene (25%), benzo(e)pyrene (24%), benzo(a)pyrene EPA (14%), benz(a)anthracene EPA (11%), dibenzo(a,h)pyrene (9%), and benzo(b)fluoranthene EPA (5%) in the total PM∑ 88 BaPeq toxicity of heavy-duty vehicle emissions ( Figure S1, project E-5). As can be seen, only three out of six listed top PAHs were EPA priority compounds, which accounted for only 31% of PM∑ 88 BaPeq levels.
In the case of biomass-burning samples (project B-1), the EPA priority PAHs contributed more to the total PM PAH toxicity (∑ 88 BaPeq) than to the heavy-duty engine emissions (E-1), on average making up to 57.5% of the total PM∑ 88 BaPeq. The dominant PAHs in biomass-burning samples were benzo(a)pyrene EPA (37%), benzo(e)pyrene (30%), benz(a)anthracene EPA (6%), benzo(b)fluoranthene EPA (5%), dibenzo(ah)anthracene EPA (4%), benzo(j)fluoranthene (3%). Overall, non-EPA PAHs were responsible for 42.5% of the PM∑ 88 BaPeq ( Figure S1, project B-1), which was still a significant contribution to the PAH toxicity level. Delgado-Saborit et al. [74] reported that the main contributor to the carcinogenic potency of PM PAH mixtures in different indoor and outdoor environments is benzo(a)pyrene, which was in agreement with our results. However, in their study, only 16 EPA PAHs were measured, and thus the carcinogenic potency of non-EPA PAHs has been overlooked. burning samples (project B-1), the EPA priority PAHs contributed more to the total PM PAH toxicity (∑88BaPeq) than to the heavy-duty engine emissions (E-1), on average making up to 57.5% of the total PM∑88BaPeq. The dominant PAHs in biomass-burning samples were benzo(a)pyreneEPA (37%), benzo(e)pyrene (30%), benz(a)anthraceneEPA (6%), benzo(b)fluorantheneEPA (5%), dibenzo(ah)anthraceneEPA (4%), benzo(j)fluoranthene (3%). Overall, non-EPA PAHs were responsible for 42.5% of the PM∑88BaPeq ( Figure S1, project B-1), which was still a significant contribution to the PAH toxicity level. Delgado-Saborit et al. [74] reported that the main contributor to the carcinogenic potency of PM PAH mixtures in different indoor and outdoor environments is benzo(a)pyrene, which was in agreement with our results. However, in their study, only 16 EPA PAHs were measured, and thus the carcinogenic potency of non-EPA PAHs has been overlooked.  Correlation between BaPeq of 88 PAHs in particle-phase (PM) and 16 EPA particle-phase PAHs analyzed for 13 projects (Table 1). Table 2. BaPeq correlations of 16 EPA priority particle-phase PAHs (PM∑16BaPeq) with a BaPeq of 88 particle-phase PAHs (PM∑88BaPeq) and BaPeq of both particle-and gas-phase PAHs. The calculations were performed using Spearman's correlation. As was mentioned above, many ambient PAH studies have been focused on characterization of particle-phase PAHs, while gas-phase PAH species have been underrepresented in air quality projects so far. It has been emphasized in the literature [35,81] that high molecular weight PAH compounds (with four or more aromatic rings) are more toxic, and thus analysis of PM, especially PM2.5, is more important than gas phase measurements. At the same time, light PAHs with 2-3 aromatic rings, such as naphthalene and naphthalene derivatives, are mainly present in the gas phase. Measurements of only PM PAHs, thus could underestimate the total PAH toxicity of air samples [60]. The contribution of gas phase PAHs to the total PAH carcinogenic potency was evaluated using two equations: the BaPeq(%)II ratio [Equation (3)] and the BaPeq(%)III ratio [Equation (4)].

Project
Equation (3) represents the contribution of the particle phase of the 16 EPA priority PAHs to their total (gas + particles) toxicity, (PM + Gas)∑16BaPeq. BaPeq(%)II ratios for all analyzed projects are summarized in Figure 3 and Table S4. The ratios varied significantly from project to project, and Figure 2. Correlation between BaPeq of 88 PAHs in particle-phase (PM) and 16 EPA particle-phase PAHs analyzed for 13 projects (Table 1). Table 2. BaPeq correlations of 16 EPA priority particle-phase PAHs (PM∑ 16 BaPeq) with a BaPeq of 88 particle-phase PAHs (PM∑88BaPeq) and BaPeq of both particle-and gas-phase PAHs. The calculations were performed using Spearman's correlation.

Project
PM∑ ∑ ∑16BaPeq/PM∑ ∑ ∑88BaPeq PM∑ ∑ ∑16BaPeq/(PM + GAS)∑ ∑ ∑88BaPeq As was mentioned above, many ambient PAH studies have been focused on characterization of particle-phase PAHs, while gas-phase PAH species have been underrepresented in air quality projects so far. It has been emphasized in the literature [35,81] that high molecular weight PAH compounds (with four or more aromatic rings) are more toxic, and thus analysis of PM, especially PM 2.5 , is more important than gas phase measurements. At the same time, light PAHs with 2-3 aromatic rings, such as naphthalene and naphthalene derivatives, are mainly present in the gas phase. Measurements of only PM PAHs, thus could underestimate the total PAH toxicity of air samples [60]. The contribution of gas phase PAHs to the total PAH carcinogenic potency was evaluated using two equations: the BaPeq(%) II ratio [Equation (3)] and the BaPeq(%) III ratio [Equation (4)].
Equation (3) represents the contribution of the particle phase of the 16 EPA priority PAHs to their total (gas + particles) toxicity, (PM + Gas)∑ 16 BaPeq. BaPeq(%) II ratios for all analyzed projects are summarized in Figure 3 and Table S4. The ratios varied significantly from project to project, and ranged between 1.8% and 84%. A dominant contribution of the 16 gas-phase EPA priority PAHs to (PM + Gas)∑ 16 BaPeq was observed for A-2, E-1, E-4, E-5, and Mi-1 projects, where the 16 EPA particle-bound PAHs were responsible only for 2.3, 2.0, 8.7, 6.1, and 1.8%, respectively. The contribution of the gas-phase PAHs to the total toxicity of the 16 EPA was also larger than that of the particle-phase 16 PAHs for most of the other projects (A-1, T-2, E-2, E-3, Mi-2)-the particle-phase 16 PAHs contributed, on average, between 28% and 38%. Only for three projects (B-1, T-1, and M-1), the 16 particle-bound EPA PAHs had a significant contribution to the (PM + Gas)∑ 16 BaPeq toxicity-above 50%. Moreover, we also found that within one project, the distribution of PAHs between gas and particle phases and, therefore, their PM BaPeq contribution, varies significantly. Our study of traffic emissions at the Fort McHenry tunnel (Baltimore, MD, USA) showed that the contribution of the particle-phase EPA PAHs to the total 16 PAH toxicity (or ∑ 16 BaPeq) is larger in winter (T-1, BaPeq(%) II = 64.8%) than in summer (T-2, BaPeq(%) II = 37.1%). During this study, summer temperatures were on average 29 • C, ranging from 25 • C to 31 • C, while in winter, temperatures were on average 0.2 • C, ranging from −9.1 • C to 13.7 • C. Gas-particle partitioning of semi-volatile compounds shifts to the gas phase at higher temperature [48,82,83], which explains the stronger contribution of gas-phase PAHs during summer months of this project. It should be also noted that the gas-particle partitioning of PAHs and other semi-volatile compounds depends not only on ambient temperature, but also on the concentration of particle-phase organics [83]. Thus, dilution of emitted PAHs and mixing with organic aerosols from other sources could change gas-particle partitioning of semi-volatile PAHs. Measurements of only particle-phase PAHs represent particle-phase toxicity only at the conditions of the measurements and cannot be used to estimate toxicity at other conditions. (PM + Gas)∑16BaPeq was observed for A-2, E-1, E-4, E-5, and Mi-1 projects, where the 16 EPA particlebound PAHs were responsible only for 2.3, 2.0, 8.7, 6.1, and 1.8%, respectively. The contribution of the gas-phase PAHs to the total toxicity of the 16 EPA was also larger than that of the particle-phase 16 PAHs for most of the other projects (A-1, T-2, E-2, E-3, Mi-2)-the particle-phase 16 PAHs contributed, on average, between 28% and 38%. Only for three projects (B-1, T-1, and M-1), the 16 particle-bound EPA PAHs had a significant contribution to the (PM + Gas)∑16BaPeq toxicity-above 50%. Moreover, we also found that within one project, the distribution of PAHs between gas and particle phases and, therefore, their PM BaPeq contribution, varies significantly. Our study of traffic emissions at the Fort McHenry tunnel (Baltimore, MD, USA) showed that the contribution of the particle-phase EPA PAHs to the total 16 PAH toxicity (or ∑16BaPeq) is larger in winter (T-1, BaPeq(%)II = 64.8%) than in summer (T-2, BaPeq(%)II = 37.1%). During this study, summer temperatures were on average 29 °C, ranging from 25 °C to 31 °C, while in winter, temperatures were on average 0.2 °C, ranging from −9.1 °C to 13.7 °C . Gas-particle partitioning of semi-volatile compounds shifts to the gas phase at higher temperature [48,82,83], which explains the stronger contribution of gas-phase PAHs during summer months of this project. It should be also noted that the gas-particle partitioning of PAHs and other semi-volatile compounds depends not only on ambient temperature, but also on the concentration of particle-phase organics [83]. Thus, dilution of emitted PAHs and mixing with organic aerosols from other sources could change gas-particle partitioning of semi-volatile PAHs. Measurements of only particle-phase PAHs represent particle-phase toxicity only at the conditions of the measurements and cannot be used to estimate toxicity at other conditions. To illustrate how much the PAH carcinogenic potency of analyzed air samples would change if 88 gas-and particle-phase PAHs are taken into account instead of only 16 particle phase EPA priority PAHs, BaPeq(%)III ratios [Equation (4)] were calculated (Table S4, Figure 4). BaPeq(%)III ratios for six projects (B-1, E-1, E-3, E-4, E-5, Mi-1) were below 10% (range: 0.2-7.4%); for four projects (A-1 T-2, E-2, Mi-2) these ratios were not higher than 20% (A-1: 14.8%; T-2: 15.6%; E-2: 10.3%; Mi-2: 15.1%); and . BaPeq ratios of 16 EPA priority particle-phase PAHs to 16 EPA priority PAHs in both gas and particle phases (BaP eq (%) II = PM∑ 16EPA BaP eq /(PM + Gas)∑ 16EPA BaP eq ); orange lines-mean values of BaP eq (%) II ; green triangles-median values of BaP eq (%) II .

Figure 5.
Top six PAHs that have highest BaPeq concentration levels in both gas and particle phases.
It is clear that the "PM-only analysis" approach could significantly underestimate the carcinogenic potency of PAH mixtures. Similar conclusions have been presented by Ramirez et al. [60]. In their work, 18 gas and particle-phase PAHs have been analyzed in air samples collected near industrial cities, and the contribution of gas-phase PAHs to the total BaPeq was found to be significant (34-86%). The 16 EPA PAHs could still be used for toxicity predictions if their BaPeq toxicity correlates well with the BaPeq toxicity of all PAH compounds in the mixture. Correlation coefficients of PM∑16BaPeq with (PM + Gas)∑88BaPeq are presented in Figure 6 and Table 2. Significantly positive correlations (r > 0.8) were observed for eight projects (A-1, A-2, B-1, T-1, T-2, M-1, E-3, E-4, E-5), suggesting that the total carcinogenic potency observed in these projects could be estimated using measurements of only the 16 EPA particle-phase PAHs. However, poor correlations were observed among similar types of PAH sources: E-1 (r = 0.054) vs. E-5 (r = 0.917). This shows that applying a correction factor to PM∑16BaPeq to estimate PM∑16BaPeq may not be applicable in all conditions and could introduce a large uncertainty in toxicity estimates. It is clear that the "PM-only analysis" approach could significantly underestimate the carcinogenic potency of PAH mixtures. Similar conclusions have been presented by Ramirez et al. [60]. In their work, 18 gas and particle-phase PAHs have been analyzed in air samples collected near industrial cities, and the contribution of gas-phase PAHs to the total BaPeq was found to be significant (34-86%). The 16 EPA PAHs could still be used for toxicity predictions if their BaPeq toxicity correlates well with the BaPeq toxicity of all PAH compounds in the mixture. Correlation coefficients of PM∑ 16 BaPeq with (PM + Gas)∑ 88 BaPeq are presented in Figure 6 and Table 2. Significantly positive correlations (r > 0.8) were observed for eight projects (A-1, A-2, B-1, T-1, T-2, M-1, E-3, E-4, E-5), suggesting that the total carcinogenic potency observed in these projects could be estimated using measurements of only the 16 EPA particle-phase PAHs. However, poor correlations were observed among similar types of PAH sources: E-1 (r = 0.054) vs. E-5 (r = 0.917). This shows that applying a correction factor to PM∑ 16 BaPeq to estimate PM∑ 16 BaPeq may not be applicable in all conditions and could introduce a large uncertainty in toxicity estimates. Figure 6. BaPeq correlation between 16 EPA particle-phase PAHs and 88 PAHs in both gas and particle phases

Conclusions
In the present study, the data of 13 different projects (650 filters and XAD cartridges, total) on the analysis of 88 gas and particle-phase PAHs were used to calculate the total BaPeq toxicity by applying a widely used TEF approach. Obtained BaPeq values were compared with those calculated for 16 EPA priority particle-phase PAHs, and we found that 16 particle-bound EPA PAHs were responsible for only 14.4% on average for the range in all projects (0.2 -41.9%) and the health risk of analyzed samples was likely to have been hugely underestimated. Although there was a significant linear relationship between PM∑16BaPeq and PM∑88BaPeq as well as between PM∑16BaPeq and (PM+Gas)∑88BaPeq for a majority of the analyzed projects, the potential carcinogenic toxicity based on BaPeq concentrations for only 16 EPA particle-phase PAHs must be interpreted carefully, especially considering possible synergistic effects of PAH mixtures. Our findings on the analysis of BaPeq values for 88 gas and particle phase individual PAH compounds showed that the contribution of non-EPA PAHs such as 1-and 2-methylnaphthlalenes, benzo(e)pyrene, dibenzo(a,i)pyrene, dibenzo(a,h)pyrene, is substantial to carcinogenic potency of analyzed air samples.
Our results stress the need for data on TEF values for additional PAHs, and the incorporation of these compounds into the health risk assessment. Measurements of gas-phase PAHs are also recommended, because of the strong contribution of the gas phase to the total BaPeq toxicity of atmospheric PAHs.

Supplementary Materials:
The following are available online at www.mdpi.com/link, Table S1: EPA priority PAHs and TEF coefficients, Table S2: TEFs assigned to 88 gas and particle phase PAHs analyzed for 9 different projects, Table S3: BaPeq calculated for samples collected for different projects, Table S4: Ratios of BaPeq calculated for samples collected for different projects, Figure S1: Top six PAHs that have highest BaPeq concentrations in particle-phase samples, Figure S2: Top six PAHs that have highest BaPeq concentrations in gas-phase samples.

Acknowledgements:
The presented projects were supported by the National Science Foundation (NSF) under Grant no. AGS1544425, NASA ROSES under Grant nos. NNX15AI48G; the Desert Research Institute's Wildland Fire Science Center (WFSC), Institute Project Assignments Internal Award (IPA, DRI), DRI's EDGES program, TetraTech Inc., the National Institute for Occupational Safety and Health (NIOSH), and the Health Effects Institute (HEI 4947-RFPA14-1/15-1). We thank Eric Fujita for supervising some of the projects, Adam Watts for providing biomass fuels to perform controlled combustion experiments, Dave Campbell, Chiranjivi Bhattarai, Figure 6. BaPeq correlation between 16 EPA particle-phase PAHs and 88 PAHs in both gas and particle phases.

Conclusions
In the present study, the data of 13 different projects (650 filters and XAD cartridges, total) on the analysis of 88 gas and particle-phase PAHs were used to calculate the total BaPeq toxicity by applying a widely used TEF approach. Obtained BaPeq values were compared with those calculated for 16 EPA priority particle-phase PAHs, and we found that 16 particle-bound EPA PAHs were responsible for only 14.4% on average for the range in all projects (0.2-41.9%) and the health risk of analyzed samples was likely to have been hugely underestimated. Although there was a significant linear relationship between PM∑16BaPeq and PM∑88BaPeq as well as between PM∑16BaPeq and (PM+Gas)∑88BaPeq for a majority of the analyzed projects, the potential carcinogenic toxicity based on BaPeq concentrations for only 16 EPA particle-phase PAHs must be interpreted carefully, especially considering possible synergistic effects of PAH mixtures. Our findings on the analysis of BaPeq values for 88 gas and particle phase individual PAH compounds showed that the contribution of non-EPA PAHs such as 1-and 2-methylnaphthlalenes, benzo(e)pyrene, dibenzo(a,i)pyrene, dibenzo(a,h)pyrene, is substantial to carcinogenic potency of analyzed air samples.
Our results stress the need for data on TEF values for additional PAHs, and the incorporation of these compounds into the health risk assessment. Measurements of gas-phase PAHs are also recommended, because of the strong contribution of the gas phase to the total BaPeq toxicity of atmospheric PAHs.

Supplementary Materials:
The following are available online at www.mdpi.com/2305-6304/5/3/17/s1, Table S1: EPA priority PAHs and TEF coefficients, Table S2: TEFs assigned to 88 gas and particle phase PAHs analyzed for 9 different projects, Table S3: BaPeq calculated for samples collected for different projects, Table S4: Ratios of BaPeq calculated for samples collected for different projects, Figure S1: Top six PAHs that have highest BaPeq concentrations in particle-phase samples, Figure S2: Top six PAHs that have highest BaPeq concentrations in gas-phase samples.
Author Contributions: V. Samburova, B. Zielinska, and A. Khlystov designed and performed experiments, field studies, and data collection. V. Samburova summarized data and wrote the paper. A. Khlystov designed the plots. A. Khlystov and B. Zielinska provided input on interpretation of results. V. Samburova and A. Khlystov revised the manuscript.

Conflicts of Interest:
The authors declare no conflicts of interest.