Currently in the United States there are no regulatory standards for ambient concentrations of Polycyclic Aromatic Hydrocarbons (PAHs), a class of organic compounds with known carcinogenic species. As such, monitoring data are not routinely collected resulting in limited exposure mapping and epidemiologic studies. This work develops the log-Mass Fraction (LMF) Bayesian Maximum Entropy (BME) geostatistical prediction method used to predict the concentration of 9 particle-bound PAHs across the US state of North Carolina. The LMF method develops a relationship between a relatively small number of collocated PAH and fine Particulate Matter (PM2.5) samples collected in 2005 and applies that relationship to a larger number of locations where PM2.5 is routinely monitored to more broadly estimate PAH concentrations across the state. Cross validation and mapping results indicate that by incorporating both PAH and PM2.5 data, the LMF BME method reduces mean squared error by 28.4% and produces more realistic spatial gradients compared to the traditional kriging approach based solely on observed PAH data. The LMF BME method efficiently creates PAH predictions in a PAH data sparse and PM2.5 data rich setting, opening the door for more expansive epidemiologic exposure assessments of ambient PAH.

Polycyclic Aromatic Hydrocarbons (PAHs) are a class of organic compounds containing 2 or more fused aromatic rings created through incomplete fuel combustion from a variety of sources including biofuel burning, wildfires, coal production, etc.^{1,2} Several species of PAHs and their metabolites have been designated by the United States Environmental Protection Agency (US EPA) as being probable human carcinogens.^{3–6} Currently the EPA only has PAH regulatory standards for drinking water and the National Institute for Occupational Safety and Health (NIOSH) has established occupational exposure limits to coal tar pitch volatiles.^{7} International organizations and other countries have established ambient concentration guidelines for one of the more toxic PAHs, benzo(a)pyrene.^{8} However, currently in the US there are no regulatory standards for ambient concentrations of PAHs. Compared to regulated ambient air pollutants, there are few epidemiologic studies that have utilized observed data or explored ambient exposures to different PAHs, which can be costly to measure.^{9,10} From a geostatistical perspective, limited ambient observed data have resulted in few studies creating maps of PAHs concentrations.^{11–14} Others have used Chemical Transport Models (CTMs) to predict PAH concentrations.^{8,15,16} However, these studies are also limited in number. As a result, there is a gap in the literature exploring ambient PAH exposures and their associations with various health endpoints. Short-term health effects include eye and skin irritation, nausea and vomiting while long-term health effects include increased risk to skin, lung and bladder cancer as well as cardiopulmonary mortality.^{7} While many of these health effects are associated with either occupational exposures or drinking water exposures, the relationship between ambient concentrations of PAH to their associated health effects has not been well explored.

Both inside and outside the US there is a lack of consistent PAH observed monitoring outside of monitoring campaigns conducted for specific studies. In contrast to the data sparse environment of PAH observed data, Particulate Matter ≤ 2.5 micrometers in diameter (PM2.5) exists in a data rich environment with a vast, consistent, historical monitoring network across the US.^{17,18} Currently there are 16 EPA designated priority PAHs, 9 of which are particle-bound.^{13} Thus, a portion of PM2.5 can be particle-bound PAH. Currently, the US state of North Carolina has no maps displaying PAH concentrations using observed data. This study explores the relationship between PM2.5 and PAH and is an extension of previous work done by Allshouse et al.^{13} that developed the log-Mass Fraction (LMF) Bayesian Maximum Entropy^{19,20} (BME) geostatistical method and applied this method to model the distribution of PAH near the World Trade Center after September 11^{th}. The observational PAH data used in that work came from the analysis of 243 PM2.5 filters from four sites spanning approximately 200 days split among the sites near and around Ground Zero set up following September 11^{th}.

In this work we analyze the PAH content of PM2.5 filters collected across the US state of North Carolina and implemented the LMF BME method to predict PAH concentration at unmonitored locations, creating the first maps of PAH across North Carolina for 2005 using observed data. Furthermore, we compare the LMF BME method with a simple Linear Regression (LR) BME method and more traditional geostatistical methods for the first time. Methods are evaluated through cross validation. Predictive maps are used to visualize the probability of exceeding PAH cutoff concentrations. Lastly, a comparison is performed between the LMF BME and other methods to learn how the relationship between PAH concentrations near wildfires may change for different prediction methods. These results provide a method for which a data sparse environment can be exploited in an efficient manner in conjunction with a data rich secondary data (e.g. PM2.5 data) environment in which the resulting relationship between the two can be applied to estimate concentrations of data sparse air pollutants elsewhere in a given domain. This cost-effective method in terms of analyzing observed data can be applied to other air pollution parameters that have not been previously mapped. This methodology opens the door for greater epidemiologic studies exploring the association between ambient concentrations of PAHs and various health endpoints.

Daily PM2.5 filters in North Carolina during 2004-2005 were collected as part of the monitoring effort needed to provide the data reported in the EPA's Air Quality Systems (AQS) data base.^{17} Of the PM2.5 filters collected during this time period, we selected 84 filters collected in 2005 and analyzed them for the following 9 species of PAHs: benz(a)anthracene, chrysene, benzo(b)fluoranthrene, benzo(k)fluoranthrene, benzo(e)pyrene, benzo(a)pyrene, indeno(1,2,3-cd)pyrene, benzo(g,h,i)perylene, dibenzo(a,h)anthracene and the summation of the 9 PAH species called Total PAH. PM2.5 has units of ^{3} and PAH has units of ^{3}.

There are approximately 8,000 space/time locations for the state of North Carolina in 2005 where daily PM2.5 is observed and recorded in AQS. PAH was estimated at these locations using surrounding PAH and PM2.5 information. There were two different PAH estimation methods: 1) a LR method consisting of a regression created from paired PM2.5 and PAH in an estimation neighborhood in which PAH is then predicted at locations where PM2.5 is known and 2) a LMF method that assumes the ratio of PAH/PM2.5 is constant within an optimized estimation neighborhood in which PAH is then predicted by applying the ratio at locations where PM2.5 is known.

Let _{h}^{th} PAH measurement is denoted as _{i}_{i}_{i}_{h}_{i}_{i}_{s}^{th} individual estimate is denoted as _{j}_{s}

The LR method is a simple linear regression of PAH with respect to PM2.5. This linear regression can be expressed at the _{i}

The subsequent sections describe how the parameters _{0} and _{1} are estimated. The relationship in _{j}

where

The LMF method was defined by Allshouse et al.^{13} as expressing the relation between PAH and PM2.5 at the _{i}

which can be rewritten as

This approach is attractive because it uses only one parameter, namely LMF, to estimate PAH based on PM2.5, making the LMF more parsimonious and more localized around the estimation location of interest. At locations _{j}_{LMF}_{,}_{j}

_{LMF}_{j}_{i}_{j}_{LMF}_{j}

The relationship in _{j}

_{0} = _{LMF,j}_{1} = 1.

The parameters _{LR}_{j}_{LMF}_{j}_{LR}_{j}_{LMF}_{j}_{LR}_{j}_{LMF}_{j}^{21}

The values found for _{LR}_{j}_{LMF}_{j}_{j}_{LR}_{j}_{LMF}_{j}

BME provides a mathematically rigorous geostatistical space/time framework for the estimation of PAH at locations where neither PAH or PM2.5 are monitored.^{19,20} BME can incorporate information from multiple data sources and is implemented using the ^{22,23} The buttress of BME has been detailed in other works,^{21,24,25} and can be summarized as performing the following steps: 1) gathering the general knowledge base (G-KB) and site-specific knowledge base (S-KB) characterizing the Space/Time Random Field (S/TRF) _{G}_{S}_{K}

In this study, we use an S/TRF to describe the variability of PAH across North Carolina in 2005. In this work _{h}_{S}_{s}_{k}_{k}

To assess the prediction accuracy of the LMF and LR methods, a Leave-One-Out Cross Validation (LOOCV) accuracy analysis is performed. For each monitoring station where observed PAH data exist, all observed data from a given station are removed one at a time and a BME prediction was conducted (without recalculating _{G}

The difference between each mean predicted PAH _{i}_{i}_{i}_{i}_{i}^{3}), Variance of Errors (VE, (^{3})^{2}), Root Mean Squared Error (RMSE, ^{3}), Mean Squared Error (MSE, (^{3})^{2}) and the squared of the Pearson correlation coefficient (^{2}, unitless) calculated between observed and mean predicted values. LMF BME and LR BME predictions were then compared to kriging (i.e. predictions created only using observed PAH data) and cokriging (i.e. predictions created using both PAH and PM2.5 observed data).

Wildfires contribute to a sizable percentage of PAH emissions in the US.^{2} The mean difference in PAH concentrations near known wildfire locations were estimated. PAH was estimated on a fine grid across North Carolina on days with observed PAH data. PAH was estimated using 4 different prediction methods: 1) kriging, 2) cokriging, 3) LR BME and 4) LMF BME. Fire data were obtained from the Federal Wildfire Fire Occurrence Website.^{26} All fires greater than or equal to one acre in North Carolina, Virginia, Tennessee and South Carolina were collected in 2005 on days for which PAH observed data were measured where the start and control date of the fires were known (

A log-transformation of PM2.5 and PAH were taken due to the skewness of observed values. For each PAH and BME estimation method (i.e. LR and LMF), the optimal values of the parameters

Across each PAH and Total PAH, the minimized MSE was consistently lower for the LMF method than the LR method with the exception of benzo(a)pyrene. With these optimized neighborhoods, estimates were created by each method and each PAH is predicted across North Carolina using BME.

The PAH estimation neighborhood for the LMF method is smaller than the LR method. Out of the previously mentioned studies, very little use observed PAH data and of those studies that do, most observed data come from short-lived monitoring campaigns. The results presented in this work utilize long-term, established PM2.5 regulatory monitoring sites. PM2.5 data is comparatively plentiful. Previously, data fusion methods have blended together multiple air pollutants that have different spatial supports.^{27,28} By developing a relationship between a few PAH observations and several PM2.5 observations, the door is opened to applying this relationship to a network with a large amount of publicly available data. Data sparse environments (e.g. PAH) can benefit from data rich secondary environments (e.g. PM2.5). However, for this relationship to be fully exploited, it must be constructed in such a manner that best utilizes the limited data set. That is, the relationship between PAH and PM2.5 must be parsimonious. The LMF method has only one parameter to be estimated, namely, LMF(_{LMP}_{j}

This work created the first maps of predicted PAH in space/time across the US state of North Carolina for 2005 using observed data. Each of the 9 PAHs was predicted on a fine grid across North Carolina every day observed PAH data were collected (41 days) in 2005 for the 4 prediction methods: kriging, cokriging, LR BME and LMF BME. Estimation method parameters can be found in

Few other studies have created maps of ambient PAH concentrations across a given area using geostatistical methods from observed data. These limited studies are due in part to the lack of observed data, much like the mapping scenario presented in this work and previous works.^{13} One previous study fit a temporal trend comparing a few long-running PAH stations from the Great Lakes region of the US and a few stations across Europe. However, only a temporal trend was fit through a regression and a spatial interpolation was not conducted.^{29} One of the few studies that created maps over a large area, displayed benzo(a)pyrene across Europe for 1990, 2001 and 2005 using a transport model.^{8} Another study creating maps of PAH across Europe utilized kriging to estimated benzo(a)pyrene for 2012 using two different chemical transport models as data.^{15} A study in Portugal used observed PAH data extracted from lichen and created maps using kriging.^{11,12} Land use regression models have also been used to estimate PAH.^{30,31} Few studies have investigated PAH bound to PM.^{32} The closest study to the LR BME method presented in this work used a monitoring campaign along with personal monitors to analyze PAH from PM2.5 in which predictions were made at unmonitored locations using kriging in Kaohsiung city, Taiwan.^{14} A regression model with a variety of explanatory variables was then applied to PM2.5 data to predict PAH. With only a handful of observed PAH data taken throughout the year, the LR BME and LMF BME method can create estimates with a corresponding uncertainty that was incorporated into the BME framework. Incorporating the PAH estimates added to the set of available data ultimately used for prediction allowing for increased spatial variation. Of the BME methods, LMF BME was superior in terms of visually distinguishing spatial variations of mean predicted PAH concentrations across the state.

A LOOCV analysis was performed across 2005 using the 4 prediction methods. Summary statistics were calculated showing performance for Total PAH (^{2} from LR BME to LMF BME. The performance statistics from kriging are similar to cokriging. This echoes the results seen in the prediction maps. Traditional incorporation of PM2.5 as a co-pollutant through cokriging adds little to the predictive captivity of PAH. Incorporating PM2.5 with the BME methods showed more substantial improvements in the cross-validation statistics, with the best performance obtained through the LMF BME method.

The LMF BME method consistently outperformed the other comparison methods as seen visually through maps and through the LOOCV statistics. Of the 4 prediction methods, kriging performed the worst. Kriging predictions were driven exclusively by the observed data. Cokriging performed similarly to kriging. Cokriging is an intuitive choice for collocated, ambient, environmental parameters in a geostatistical setting. In the literature, to the best of our knowledge, cokriging has not been used to predict ambient PAH concentrations, making it an ideal candidate method to explore. In this work the cokriging cross-covariance is able to capture the relationship between PAH and PM2.5. However, as seen through predictive maps and through cross validation, the cokriging incorporation of PM2.5 contributes little in terms of predictive capacity. Linear regression is another intuitive choice with collocated data. The LR BME method shows a marked improvement visually and through estimation accuracy. The LR method is able to estimate PAH at PM2.5 space/time locations using an optimized neighborhood customized for each PAH. However, LR performed consistently worse than the LMF method. The LR method uses 2 parameters (i.e. _{0} and _{1} while the LMF method uses only one. We hypothesize that this difference in the number of parameters influences cross validation performance.

In a geostatistical framework, predictions come in the form of a PDF with a corresponding mean and variance. With this PDF, the probability of exceeding a given value can be calculated. An annual benzo(a)pyrene concentration of 0.25 ^{3} has been suggested in the United Kingdom.^{8} With this standard in mind, the probability of exceeding this cutoff was calculated for annual benzo(a)pyrene concentrations on a fine grid in North Carolina in 2005 for each prediction method by taking the mean and variance of daily benzo(a) pyrene predictions (^{3} cutoff in turn decreases from kriging to LMF BME. Across methods, the region of the state with the relatively highest probability of exceedance is maintained as Western North Carolina as well as the border with the US state of South Carolina. Across all prediction methods, the probability of exceedance remained relatively low with the maximum probability of exceedance remaining below 0.50. The area covered from increasing probabilities of exceedance increases across the 4 prediction methods (^{2} having a probability of exceedance ≥ 0.15. We see the BME methods were better able to differentiate areas of high and low probabilities of exceedance. The LMF BME was best able to distinguish the maximum probability of exceedance. Neither the kriging or cokriging maps contain any area with a probability of exceedance ≥ 0.30. The LMF BME method has 2.5 times the area with ≥ 0.30 probability of exceedance compared to the LR BME method (i.e. 6,480 ^{2} and 2,592 ^{2}, respectively). Through having more realistic ambient predictive gradients, the LMF BME method becomes an effective tool to identify areas of exceedance of different PAH concentrations.

The mean difference in PAH predictions (for the 9 PAHs and Total PAH) as calculated through the 4 prediction methods was found through a two-sampled

This work investigates ambient concentrations of a set of particle-bound PAHs. Ambient concentrations alone cannot distinguish sources. However, there are PAH ratios associated with certain sources. The diagnostic ratio of indeno(1,2,3-c,d)pyrene/(indeno(1,2,3-c,d)pyrene + benzo(g,h,i)perylene)=0.62 is associated with wood burning.^{8} This ratio was calculated for March 5, 2005 data across all 4 prediction methods (

Gathering information about wildfire smoke has become increasingly important as the number of large wildfires have increased in recent years.^{33} The chronic health effects of wildfire smoke for firefighters and the general population is currently lacking or sparse in the literature.^{34–36} The LMF BME method was better able to distinguish higher significant differences in PAH concentrations near known fire locations compared with other prediction methods. Of the 4 prediction methods the LMF was the only method that showed statistically significant, positive differences around areas with fires across all 9 PAHs and Total PAH. Although each fire may have a different acreage burned and the same buffer size was used for all the fires, the significance implies an association. Depending on the acreage burned from a fire, the type of vegetation burned and the duration of the fire, the smoke produced may be long lasting and may have long range transport. Smoke may have lingering effects past the control date of a fire. When the control date of fires is extended by one day, the kriging and cokriging methods have more PAHs with a statistically significant increase in concentrations near versus far from fires (

The LMF BME method allows for straightforward predictions of PAHs to be used for exposure assessments. There are a plethora of studies exploring the association between ambient PM2.5 and various health endpoints.^{37–39} However, there are far less studies that explore ambient PAH exposures and associated health effects. Occupational inhalation exposures and associated health outcomes including lung cancer have been more thoroughly investigated in comparison to ambient exposures.^{7} Few studies have investigated chronic ambient concentrations of PAHs. Many of the epidemiologic studies that have been explored investigate respiratory illnesses such as lung cancer and pulmonary function.^{7,15,40} However, these studies are small. The lack of long-term ambient concentrations to PAHs may be related to inadequate exposure data. Analyzing PM filters for specific PAHs can be very costly, making it difficult of obtain larger amounts of observed data needed for exposure assessment.^{9} The LMF BME method allows for an efficient and cost-effective way to utilize minimal PAHs observed data. The LMF BME method can be easily utilized to fill in this clear gap in the literature. Tied with corresponding health data, ambient predictions calculated through the LMF BME method could be used to assign exposure. Health metrics can then be calculated from the exposures. This opens the door to investigate possible health endpoints as well as assigning risk.

This work created the first maps of ambient PAH concentration across the US state of North Carolina using observed data through the LMF BME geostatistical method. This method developed a relationship between paired PAH and PM2.5 data in a manner that is a parsimonious and cost-effective that can be utilized in a data sparse environment. The LMF BME method outperforms more traditionally used geostatistical methods and has the ability to elucidate a significant association between PAH predictions and known fire locations. The LMF BME method has the potential to be used to assign exposure in epidemiologic analyses to fill in the significant knowledge gap currently existing in the literature between ambient PAH exposures and potential health outcomes.

^{3})^{2}).

_{PAH}^{3})^{2}, _{PM}^{3})^{2} and _{PAH},_{PM}^{3}) * (^{3}).

^{3}. *mean difference is statistically significant (p-value≤ 0.05), ^{#}mean difference > 0.

^{3}. *mean difference is statistically significant (p-value≤ 0.05), ^{#}mean difference > 0.

^{3}. *mean difference is statistically significant (p-value≤ 0.05), ^{#}mean difference > 0.

^{3}. *mean difference is statistically significant (p-value≤ 0.05), ^{#}mean difference > 0.

This research was supported in part by the National Institute on Aging (NIA) under award number R01AG033078, the National Institute of Occupational Safety and Health (NIOSH) under grant 2T42/OH-008673, the National Institute of Environmental Health Sciences (NIEHS) under grant T32ES007018 and an appointment to the Internship/Research Participation Program at the Office of Research and Development, U.S. Environmental Protection Agency, administered by the Oak Ridge Institute for Science and Education through an interagency agreement between the U.S. Department of Energy and EPA. The views expressed in this document are solely those of the authors and do not necessarily reflect those of the NIA, NIOSH, NIEHS or the EPA. This article was reviewed in accordance with the policy of the National Exposure Research Laboratory, U.S. Environmental Protection Agency, and approved for publication.

Map of benzo(b)fluoranthrene (^{3}). Maps of mean benzo(b)fluoranthrene concentration for North Carolina on April 16, 2005 across the 4 prediction methods: (a) kriging, (b) cokriging, (c) Linear Regression BME, (d) log-Mass Fraction BME. Square markers indicate observed data, circle markers indicate PAH estimates, X's mark known fires for that day with a 100

Probability of exceedance. Probability of annual benzo(a)pyrene exceeding 0.25 ^{3} across North Carolina in 2005 as predicted by (a) kriging, (b) cokriging, (c) Linear Regression BME and (d) log-Mass Fraction BME.

PAH ratios. Ratio of indeno(1,2,3-c,d)pyrene/(indeno(1,2,3-c,d)pyrene+benzo(g,h,i)perylene) on March 5, 2005 in North Carolina across the 4 prediction methods: (a) kriging, (b) cokriging, (c) Linear Regression BME, (d) log-Mass Fraction BME. Square markers indicate the ratio of observed data, circle markers indicate the ratio of PAH estimates, X's mark known fires for that day with a 100

Neighborhood optimization for each PAH estimate. Optimized

Linear Regression | log-Mass Fraction | |||||
---|---|---|---|---|---|---|

PAH | S/T Metric (km/days) | MSE (^{3})^{2} | S/T Metric (km/days) | MSE (^{3})^{2} | ||

benz(a)anthracene | 14 | 0.891 | 1.128 | 5 | 0.839 | |

chrysene | 7 | 0.600 | 0.979 | 5 | 0.839 | |

benzo(b)fluoranthrene | 7 | 0.863 | 1.358 | 5 | 0.899 | |

benzo(k)fluoranthrene | 14 | 0.895 | 1.375 | 5 | 0.842 | |

benzo(e)pyrene | 14 | 0.895 | 1.006 | 2 | 0.868 | |

benzo(a)pyrene | 14 | 0.895 | 5 | 0.899 | 1.417 | |

indeno(1,2,3-c,d)pyrene | 14 | 0.891 | 0.892 | 2 | 0.868 | |

benzo(g,h,i)perylene | 14 | 0.895 | 0.757 | 2 | 0.777 | |

dibenzo(a,h)anthracene | 14 | 0.772 | 1.532 | 3 | 0.820 | |

Total PAH | 14 | 0.895 | 0.890 | 3 | 0.820 |

Cross validation statistics. Leave-One-Out Cross Validation statistics for Total PAH (summation of the 9 PAHs) comparing observed and predicted concentrations across the 4 prediction methods for North Carolina in 2005. ME is Mean Error, VE is Variance of Error, RMSE is Root Mean Squared Error, MSE is Mean Squared Error and ^{2} is the Pearson correlation coefficient squared.

Statistic | Kriging | Cokriging | Linear Regression BME | log-Mass Fraction BME |
---|---|---|---|---|

ME (^{3}) | -0.145 | -0.137 | -0.102 | -0.042 |

VE (^{3})^{2} | 0.806 | 0.782 | 0.764 | 0.591 |

RMSE (^{3}) | 0.904 | 0.890 | 0.875 | 0.765 |

MSE (^{3})^{2} | 0.818 | 0.792 | 0.766 | 0.586 |

^{2} (unitless) | 0.747 | 0.752 | 0.744 | 0.821 |

Area (in ^{2}) covered corresponding to increasing probabilities of exceeding the average annual benzo(a)pyrene standard of 0.25 ^{3} among the prediction locations in and around North Carolina in 2005.

Probability of Exceedance | ≥0.10 | ≥0.15 | ≥0.20 | ≥0.25 | ≥0.30 |
---|---|---|---|---|---|

Kriging | 162,000 | 119,232 | 5,184 | 0 | 0 |

Cokriging | 143,856 | 54,432 | 0 | 0 | 0 |

Linear Regression BME | 164,592 | 139,968 | 36,288 | 10,368 | 2,592 |

log-Mass Fraction BME | 156,816 | 116,640 | 28,512 | 15,552 | 6,480 |

Mean difference in PAH near versus far from fires. 95% confidence intervals comparing the mean difference in predicted PAH near (within 100 ^{3}.

PAH | Kriging | Cokriging | Linear Regression BME | log-Mass Fraction BME |
---|---|---|---|---|

benz(a)anthracene | (-4.94E-03,-2.17E-03) | (-2.61E-03,-1.07E-05) | (-1.26E-03,1.07E-03) | (1.57E-03,4.38E-03)^{,} |

chrysene | (-6.67E-03,-3.53E-03) | (-3.74E-03,-8.28E-04) | (-9.40E-04,1.67E-03) | (2.07E-03,5.39E-03)^{,} |

benzo(b)fluoranthrene | (3.98E-03,1.11E-02)^{,} | (3.78E-03,1.10E-02)^{,} | (1.09E-02,2.23E-02)^{,} | (2.36E-02,3.02E-02)^{,} |

benzo(k)fluoranthrene | (3.14E-03,6.47E-03)^{,} | (2.32E-03,5.01E-03)^{,} | (5.27E-03,7.94E-03)^{,} | (7.80E-03,1.08E-02)^{,} |

benzo(e)pyrene | (-2.92E-03,2.23E-03) | (-3.17E-03,1.71E-03) | (5.22E-03,9.80E-03)^{,} | (1.83E-02,2.49E-02)^{,} |

benzo(a)pyrene | (-3.83E-03,1.84E-03) | (-6.24E-03,-8.42E-04) | (2.23E-03,1.37E-02)^{,} | (5.14E-03,1.02E-02)^{,} |

indeno(1,2,3-c,d)pyrene | (1.87E-02,3.05E-02)^{,} | (1.73E-02,2.87E-02)^{,} | (2.04E-02,3.24E-02)^{,} | (4.79E-02,6.11E-02)^{,} |

benzo(g,h,i)perylene | (3.04E-02,4.27E-02)^{,} | (2.54E-02,3.66E-02)^{,} | (2.72E-02,4.06E-02)^{,} | (3.13E-02,4.06E-02)^{,} |

dibenzo(a,h)anthracene | (-2.07E-02,-1.36E-02) | (-1.71E-02,-1.07E-02) | (-4.80E-03,2.16E-03) | (1.90E-03,8.77E-03)^{,} |

Total PAH | (2.28E-02,6.75E-02)^{,} | (2.13E-02,6.57E-02)^{,} | (6.29E-02,1.03E-01)^{,} | (1.72E-01,2.30E-01)^{,} |

mean difference is statistically significant (p-value≤ 0.05),

mean difference > 0.