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. 2024 Dec 17;58(52):23117–23126. doi: 10.1021/acs.est.4c09139

Assessing Wildfire Impact on Diffusive Flux of Parent and Alkylated PAHs: A Pilot Study of Soil–Air Chemical Movement before, during, and after Wildfires

Kelly E O’Malley 1, Christine C Ghetu 1, Diana Rohlman 2, Kim A Anderson 1,*
PMCID: PMC11697327  PMID: 39689904

Abstract

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The global wildfire risk is predicted to rise due to contributing factors of historical fire management strategies and increases in extreme weather conditions. Thus, there is a need to better understand contaminant movement and human exposure to wildfire smoke. Vapor-phase polycyclic aromatic hydrocarbons (PAHs) are elevated during wildfires, but little is known about how these chemicals move during and after wildfire events for exposure risk assessment. Paired air and soil pore air passive samplers were deployed before, during, and after wildfires to determine diffusive flux of vapor-phase parent (p-PAH) and alkylated (a-PAH) PAHs in the Western United States. Naphthalene and 2-methylnaphthalene contributed to most of the volatilization and deposition (6.3–89%) before and after a wildfire. Retene (41%) and phenanthrene (27%) contributed substantially to deposition during a wildfire. During wildfires, the number of PAHs in deposition increased at sites with worse air quality. Most p-PAHs and a-PAHs were either depositing or near equilibrium after a wildfire, except for retene at several locations. A majority (≥50%) of PAHs had a 50% magnitude difference between flux before and after a wildfire. This study increases the understanding of PAH movement and exposure during each stage of the wildfire cycle.

Keywords: passive sampling, community-engaged, disaster research, wildfire smoke, polycyclic aromatic hydrocarbons

Short abstract

Passive samplers were deployed before, during, and after wildfires to evaluate diffusive flux between soil and air. PAH flux differed between pre-wildfire and post-wildfire. PAH deposition increased as air quality decreased.

Introduction

Wildfires have become an emerging air quality concern for the public due to the increased number of large wildfires in recent years.1 Past fire management and increases in extreme weather conditions such as wind, higher temperatures, and lower humidity have led to environments more conducive to wildfire occurrence and spread.2,3 Therefore, there is a need to better understand human exposure to wildfire smoke.

Wildfires can emit a variety of pollutants ranging from gases, metals, particulate matter, and organic chemicals.4 Polycyclic aromatic hydrocarbons (PAHs) have been detected in both the vapor (gas) phase and bound to particulates in previous wildfire studies.5−9 PAHs are a class of organic compounds consisting of unsubstituted parent (p-PAHs) and substituted constituents, such as alkylated PAHs (a-PAHs) that can be formed through incomplete combustion of organic materials during wildfires. Exposure to PAHs has been associated with reproductive effects, respiratory issues, cardiotoxicity, immunotoxicity, neurotoxicity, and various cancers.10−19 Alkylated PAHs are more abundant in the environment and can be more toxic than their parent constituent in some instances.20 PAH risk assessment studies have primarily focused on particulate-bound PAHs, but both particulate-bound and vapor-phase PAHs contribute to toxicity.21,22 Specifically, previous work indicates that vapor-phase PAHs contribute between 34 and 86% of the carcinogenic potency from inhalation risk. Therefore, analyzing PAHs in the vapor-phase is important for risk assessments.22

In the environment, air-soil pollutant exchange processes occur constantly to reach equilibrium in response to changes like temperature or soil microbial activities.23−25 The main mechanisms of exchange are through atmospheric (dry or wet) deposition to the soil and re-emission, where the soil is a potential sink and source for contaminants. However, little is understood about how these processes may be enhanced or significantly changed (i.e., volatilization to deposition) during wildfires. Elevated PAH levels have been reported in vegetation and soil post-wildfire.26,27 PAHs may volatilize from the soil to reach equilibrium, resulting in additional exposure after a wildfire when visible smoke has dissipated. It is also unknown if there are long-term changes in PAH movement compared to conditions prewildfire. Additionally, research on soil-air exchange of PAHs has focused mainly on seasonal or site (urban, rural, or industrial) differences.28−37

With the occurrence of wildfire size increasing in the Pacific Northwest, there are many community concerns related to chemical exposures from wildfire smoke. Informal conversations with community members and the Oregon State University (OSU) Extension service in 2017 led to characterizing deposition of wildfire smoke in agricultural environments as a primary community concern.38 Wildfire studies can be facilitated by working with trained community members through community engaged research (CEnR) where participants can quickly respond to and sample during a wildfire event. Passive sampling devices (PSDs) are a beneficial tool for CEnR due to low maintenance, ease of use, and ability to be mailed at ambient temperature. PSDs are capable of sequestering vapor-phase organic compounds in a nonselective manner by mimicking biological membranes.

The objective of this pilot study was to assess the impact of wildfires on diffusive flux of vapor-phase p-PAHs and a-PAHs. We hypothesized that p-PAHs and a-PAHs would (1) deposit more (number and magnitude) to soils with increasing Air Quality Index (AQI) values during wildfires, (2) volatilize after a wildfire, and (3) that there would be a difference in flux before and after a wildfire. To our knowledge this is the first study to look at vapor-phase PAH soil-air exchange before, during, and after wildfire smoke events across three states in the Western United States (U.S.).

Materials and Methods

Sample Locations

Sampling spanned approximately 966 km (∼600 mi) across the Western U.S. from 2018 to 2020 at seven locations in Washington, Oregon, and California (Table S1). The locations are in six different level IV ecoregions with distinct climates, geographies, and vegetation (Table S2). Locations were selected through convenience sampling using a CEnR approach. Recruitment and engagement with participants were conducted under the Institutional Review Board from Oregon State University (protocol # IRB-2019–0312) and are described in Ghetu et al. 2022.5 Briefly, the study coordinator contacted prospective participants (18 or older) residing in wildfire high-risk areas and informed them of the details of the study. Eligible participants received a sampling kit with training materials via mail (see SI). Cascade Locks, OR and Carson, WA locations were combined into “Columbia Gorge” due to participant relocation within a 16 km area between sampling years (Figure S1 and Table S1). Samples collected during wildfires captured the smoke impact from fires within nine level IV ecoregions as described in SI and Table S3.

Sample Preparation, Deployment, and Processing

Samples were collected using nonadditive low-density polyethylene (LDPE), prepared as previously described.39,40 Briefly, LDPE from Brentwood Plastics Inc. was cut into 1.1 m strips and conditioned with a hexane extraction occurring every 24 h for 3 consecutive days. Individual strips were then spiked with three performance reference compounds (PRCs), fluorene-d10, pyrene-d10, and benzo[b]fluoranthene-d12, to calculate the sampling rate of analytes using PRC loss (Table S4). LDPE strips were stored in sealed polytetrafluoroethylene (PTFE) bags to maintain analyte stability during transport.41

Trained participants deployed a T-shaped air cage approximately 1 m above the ground (representing the adult breathing zone) along with a soil pore air (hereafter referred to as soil) box. The soil box was placed on top of the soil immediately adjacent to the air cage at each location (Figure S2).30,42 Each air cage and soil box contained 5 PRC infused LDPE strips. Samplers were deployed between 10 and 48 days before, during and after a wildfire. Ambient temperature and relative humidity were measured using a temperature logger.

After deployment, participants mailed the sampling kits back to the Food Safety and Environmental Stewardship lab at OSU. The LDPE was cleaned using isopropanol to remove particulates and extracted with hexanes.40,41,43 The dialysate was then reduced to 1 mL with nitrogen.43 Deuterated PAH surrogates (naphthalene-d8, acenaphthylene-d8, phenanthrene-d10, fluoranthene-d10, chrysene-d12, benzo[a]pyrene-d12, and benzo[ghi]perylene-d12) were added prior to extraction to account for chemical recovery (Table S5). Internal standard (perylene-d12) was added to each composite sample following extraction and stored at −20 °C prior to analysis.

Instrument Analysis

Following extraction, samples were analyzed for 63 PAHs (41 p-PAHs and 22 a-PAHs) with a paired Agilent 7890B gas chromatograph (GC) and 7000C triple-quadrupole mass spectrometer (MS/MS) using an Agilent PAH-select column (30 m x 250 μm × 0.15 μm). Instrument parameters and detection and quantitative limits are detailed in Anderson et al. 2015 and Table S6–S8.44 Automatic surrogate correction was used to quantitatively account for losses through the extraction process. Calibration curves had a minimum of 5 points with an r2 ≥ 0.99. This method uses MassHunter Quantitative Analysis v.B.06.00 SP1 build 6.0.388.1 (Agilent Corp. Wilmington, DE) software.

Quality Control

Quality control (QC) samples accounted for approximately 70% of all samples analyzed. Continuous calibration verifications (CCVs), containing all analytes in the method, were run prior to and at the end of each analysis to ensure data quality objectives were met. A CCV was considered passing if target analytes were within ±30% of the true value for at least 80% of the target analytes. Instrument blanks were also run at the beginning of the analysis to ensure there was no contamination within the system. Construction, trip, lab processing, cleaning, and reagent blanks were used for each sampling year. Duplicates and over spikes were also used to ensure data quality. Analytes detected in the lab processing blanks were background subtracted from samples as described in Table S9. Air cages were deployed in triplicate at Newport, OR.

Soil–Air Flux Calculations

Diffusive flux between soil air and air was calculated as described in Donald and Anderson 2017.30 Briefly, Fick’s law of diffusion was used to calculate flux:45−47

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where a positive value (ng/m2/d) indicates volatilization from soil to air and a negative value indicates deposition from air to soil. Csoil and Cair are environmental concentrations of soil and air (ng/m3), and their calculations are described in SI and provided in Tables S20–S22. The boundary layer (δL, m) is 0.001, and DT (m2/d) is the temperature-corrected diffusivity in air where pyrene at 298 K is used as the reference for DT.45,48−50 DT calculations for each target analyte are described and provided in Tables S23–S25. Diffusive flux values were only calculated when both paired air and soil concentrations were above the limit of detection (LOD).

Three air cages were co-deployed with one soil box at Newport, OR in 2018 and 2019 to evaluate the uncertainty of the measured flux values. Donald and Anderson 2017 evaluated the between-box relative standard deviation (RSD) for air or soil and found a significant difference for soil.30 This is likely due to the heterogeneous nature of soil compared to air, so only one soil box was deployed to represent a composite of the soil conditions at Newport, OR.

RSDs for PAH flux in 2018 and 2019 were calculated each year for compounds detected in both air and soil matrices as well as all triplicates (Table S17–S18). Average percent RSDs were then calculated from the 2018 and 2019 triplicates and applied to the corresponding compounds detected across all the samples (Table S19). Sum PAH RSDs were averaged between 2018 and 2019 and were applied to PAHs not detected in the Newport, OR triplicates.

Defining Wildfire Status

For each sampling campaign, a location was categorized as being “Pre-Wildfire”, “Wildfire” and “Post-Wildfire”. To determine the category of each location during sampling and the duration of a smoke event, we utilized AQI and NOAA Hazard Mapping System (HMS) smoke density tools.51 AQI describes the level of outdoor air pollution and associated health concerns and HMS provides daily smoke conditions of an area. PAH concentrations in LDPE are time-integrated throughout the deployment period, reflecting time-weighted averages of the environmental concentrations. Verified Air Quality System AQI data for PM2.5 was collected and averaged from geographically representative EPA stationary air monitors within 97 km (∼60 miles) of the sampling location (Table S10). To ensure a minimum signal above ambient background, we defined the “Wildfire” status as an event having an AQI above 100 and “heavy” smoke density for a minimum of three consecutive days (n = 4) (Table S11–S12). “Post-Wildfire” is defined as a smoke event that occurred within nominally a month prior to sampling (n = 7) (Table S13). “Pre-Wildfire” is defined as no smoke event occurring for one year prior to sampling (n = 7) (Tables S14–S15).

Differences in PAH Flux between Pre-Wildfire and Post-Wildfire

Due to small sample sizes, a formal statistical analysis could not be used to assess differences in flux Pre- and Post-Wildfire. Instead, we defined criteria to compare flux. Noteworthy differences in PAH flux between Pre-Wildfire and Post-Wildfire were defined by meeting any of the following criteria: (1) there was a 50% higher level of flux during one wildfire condition compared to the other at the same location, (2) there was a directionality change (ex: volatilization to deposition) between Pre-Wildfire and Post-Wildfire, or (3) a PAH was below detection limits Pre-Wildfire but detected Post-Wildfire. We considered there to be enough evidence to suggest a difference between Pre-Wildfire and Post-Wildfire events if ≥50% of the total PAHs met any of the criteria.

Results and Discussion

Across the three years and seven locations, 51 of the 63 targeted PAHs exhibited diffusive flux in at least one location. The most frequent PAHs in flux, that were also detected in all samples across all time points, were naphthalene, 1-methylnaphthalene, 2-methylnaphthalene, phenanthrene, 2-methylphenanthrene, dibenzothiophene, retene, and fluoranthene. Detected in over 75% of samples, fluorene (94%), pyrene (94%), 2,6-dimethylnaphthalene (89%), and 1,4-dimethylnaphthalene (78%), 1,5-dimethylnaphthalene (78%) and 1,6- and 1,3-dimethylnaphthalene (78%) were also frequently in diffusive flux. Most of these PAHs are of lower molecular weight (LMW PAHs) as defined by being with 2 or 3 rings. LMW PAHs are more abundant in the vapor-phase, which may explain why their movements are more influenced by wildfires than higher molecular weight (HMW) PAHs (4 or more rings).52,53

Diffusive flux was measured for 28 PAHs throughout the Pre-Wildfire samples. Sum PAH flux of Pre-Wildfire samples was 210,000 ng/m2/d (volatilization) and −36,000 ng/m2/d (deposition) (Figure S4). Naphthalene and 2-methylnaphthalene contributed to 89% and 6.3% of the Pre-Wildfire volatilization. Naphthalene, 2-methylnaphthalene, 1-methylnaphthalene, and acenaphthene contributed to 65%, 9.9%, 7.6%, and 6.6% of the Pre-Wildfire deposition, respectively.

Diffusive flux was measured for 52 PAHs throughout the Wildfire samples. Sum PAH flux of Wildfire samples was 250,000 ng/m2/d (volatilization) and −44,000 ng/m2/d (deposition) (Figure 1). Naphthalene, 2-methylnaphthalene, and 1-methylnaphthalene contributed to 72%, 12%, and 7.1% of the Wildfire volatilization, respectively. Wildfire deposition had a notably different composition of major contributors that included retene, phenanthrene, 1-methylphenanthrene, pyrene, and 2-methylphenanthrene at 41%, 27%, 9.1%, 7.4%, and 5.7%, respectively.

Figure 1.

Figure 1

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Individual PAH volatilization and deposition contribution of ≥5% of the sum flux for all Wildfire samples. All PAHs contributing to less than 5% are categorized as “Other”. The number of PAHs in the “Other” category is indicated in parentheses. Pre-Wildfire and Post-Wildfire volatilization and deposition are shown in Figure S4.

Diffusive flux was measured for 27 PAHs throughout the Post-Wildfire samples. Sum PAH flux of Post-Wildfire samples was 91,000 ng/m2/d (volatilization) and −56,000 ng/m2/d (deposition) (Figure S4). Naphthalene and 2-methylnaphthalene contributed to 85% and 8.3% of the Post-Wildfire volatilization, respectively. Naphthalene, 2-methylnaphthalene, 1-methylnaphthalene, and acenaphthene contributed to 45%, 21%, 13%, and 7.4% of the Post-Wildfire deposition, respectively.

Wildfire PAH Diffusive Flux

P-PAH Diffusive Flux

Four samples at different sites in 2020 were characterized as Wildfire: Newport, OR, Prineville, OR, Corvallis, OR, and St. Helena, CA (Table S1). Naphthalene, phenanthrene, fluorene, acenaphthene, anthracene, and pyrene had the greatest magnitude of flux during a wildfire (50 to 67,000 ng/m2/d and −32 to −11,000 ng/m2/d). However, individual PAHs varied in directionality of flux at the different sites (Figure 2). Notably, naphthalene had the greatest magnitude of flux and was volatilizing at all four sites (Figure S5). Phenanthrene was volatilizing at Newport, OR (3,300 ng/m2/d) but was depositing at Prineville, OR (−470 ng/m2/d), Corvallis, OR (−11,000 ng/m2/d), and St. Helena, CA (−180 ng/m2/d). Fluorene was volatilizing at Newport, OR (1,000 ng/m2/d) and Prineville, OR (560 ng/m2/d) but depositing at St. Helena, CA (−400 ng/m2/d) and Corvallis, OR (−330 ng/m2/d). Acenaphthene was volatilizing at Newport, OR (50 ng/m2/d), Corvallis, OR (1,900 ng/m2/d) and St. Helena, CA (540 ng/m2/d) and below the LOD at Prineville, OR. These results are discussed further in the “Wildfire Comparative Analysis” section.

Figure 2.

Figure 2

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Individual PAH Wildfire flux for p-PAHs (top row) and a-PAHs (bottom row) for Newport, OR, St. Helena, CA, Prineville, OR, and Corvallis, OR. NOTES: Red box around Corvallis, OR indicates different scales. Naphthalene and alkylated naphthalenes are not included due to larger order of magnitudes and are in Figure S5.

A-PAHs Diffusive Flux

A-PAHs were mostly in deposition during wildfires. The a-PAHs with the greatest magnitude of flux during wildfires were 1-methylnaphthalene, 2-methylnaphthalene, 1-methyl-7-isopropyl phenanthrene (retene), 2-methylphenanthrene, and 1-methylphenanthrene (120 to 13,000 and −32 to −17,000 ng/m2/d) (Figure 2). These a-PAHs primarily deposited to the soil during wildfire smoke events at Prineville, OR, Corvallis, OR, and St. Helena, CA. At Newport, OR, only retene (−450 ng/m2/d) was in deposition while 2-methylphenanthrene (280 ng/m2/d), and 1-methylphenanthrene (120 ng/m2/d) were in volatilization. A-PAHs in flux during a wildfire tended to move in the same direction as the parent constituent, except for 1-methylpyrene and retene at Newport, OR and 3,6-dimethylphenanthrene at St. Helena, CA (Table S27).

Wildfire Comparative Analysis of PAH Soil–Air Exchange in Other Environments

There are no previous studies that measured diffusive flux of PAHs between soil and air during wildfires. Instead, diffusive flux of PAHs from this study are compared to studies where soil-air diffusive flux of vapor-phase PAHs was measured in other environment types and seasons. These values are listed in Table S29. Generally, PAHs were volatilizing at semiurban Bursa, Turkey, mixed urban environments at Kathmandu and Pokhara, Nepal, and urban green space located in the city center of Shanghai, China, and industrial, agricultural, residential, and rural areas throughout the Pearl River Delta (PRD) region in South China, while PAHs were depositing in the present study.54−57 For instance at most sites, phenanthrene, pyrene and fluoranthene had a greater magnitude of deposition during a wildfire compared to Bursa, Kathmandu and Pokhara, Shanghai, and the PRD.54−57 Corvallis, OR had the highest impact from wildfire smoke, and it was the only wildfire site that generally had more deposition of anthracene, benz[a]anthracene, chrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[a]pyrene, indeno[1,2,3-cd]pyrene, and benzo[ghi]perylene than sites in Bursa, Kathmandu and Pokhara, the PRD, and Shanghai (Table S29).54−57 Of these PAHs, all had greater deposition at suburban Izmir, Turkey than Corvallis, OR, except benzo[a]pyrene, a known carcinogen.29 Additionally, indeno[1,2,3-cd]pyrene and benzo[ghi]perylene deposition at Corvallis, OR was comparable to some industrial, forestry, and agricultural regions in the PRD where combustion of coal and refined petroleum and oil spill are the primary sources of PAHs.57 In contrast, Izmir had greater magnitudes of deposition of phenanthrene, fluorene, fluoranthene, and pyrene than all wildfire sites.29 Sampling at Izmir occurred on campus at Dokuz Eylul University 4m above a young conifer forest. Sources contributing to deposition may include residential areas and a highway located 2 km and 0.5 km away from motor vehicle emissions.29

Overall, wildfires can temporarily lead to PAH movement opposite, and even sometimes more dramatic than some urban, suburban, and industrial environments globally. We observed general trends of PAH deposition during wildfires in the western U.S. that far exceeded other rates, regardless of residential, agricultural, industrial, and traffic sources. The only exception was in Izmir, Turkey where a nearby highway may have contributed significantly to the PAH deposition that surpassed rates greater than the wildfires. Previous vapor-phase PAH diffusive flux studies primarily measured PAHs on the U.S Environmental Protection Agency’s priority PAH list consisting of only p-PAHs. Here, we characterized both p-PAHs and a-PAHs during wildfire events, given the association with health concerns.58−64

Wildfire PAH Flux and AQI

Sum vapor-phase PAH concentrations in outdoor air have been shown to increase at higher AQI smoke impacted sites.5 These higher PAH concentrations in the air would subsequently suggest deposition could also increase with AQI. We posited that vapor-phase PAHs undergo deposition during smoke events with a high AQI (e.g., greater smoke density). To address this, we looked for a positive trend between the fraction of p-PAHs and a-PAHs in flux that were in deposition and the average AQI during the Wildfire deployment (Figure 3). P-PAHs and a-PAHs had an r2 of 0.705 and 0.812 with AQI during the wildfire smoke events, indicating a general trend for deposition of both p-PAHs and a-PAHs irrespective of smoke source from the different ecoregions. However, when looking at individual PAHs or specific sites, not all PAHs follow this trend. Naphthalene and its alkylated constituents were always in volatilization during wildfire smoke events, whereas most of the other PAHs were in deposition. Volatilization of naphthalenes during a wildfire may be attributable to their lower molecular weights, lower boiling points, and abundance in the environment. This may lead to more soil-air exchange of naphthalenes at ambient temperatures to reach equilibrium. Interestingly, at Newport, OR, which had the lowest AQI, most p-PAHs and a-PAHs were not in deposition. Notably, this site barely reached our threshold criteria for wildfire. Wildfire smoke at Newport, OR may have been below a certain severity (e.g., AQI) where there would be a shift from volatilization to deposition. Additionally, the difference in results could be due to the soil type at Newport, OR compared to the other sites (Bandon fine sandy loam, versus silt, gravel, or clay loams (Table S30)). Generally, sandier soils have less carbon content, so PAHs have less tendency to remain in the soil and will volatilize into the air.65−67

Figure 3.

Figure 3

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Average AQI at sites during a Wildfire and the corresponding proportion of A) p-PAHs and B) a-PAHs in deposition from air to soil.

Post-Wildfire PAH Diffusive Flux

P-PAHs

Seven samples over three years at five different sites were characterized as Post-Wildfire: Newport, OR (2018), Prineville, OR (2018), Corvallis, OR (2018 and 2020), Cobb, CA (2018), and Columbia Gorge (2018 and 2020). Most p-PAHs were either depositing or near equilibrium, except for acenaphthene (450 ng/m2/d) and fluoranthene (230 ng/m2/d) at Newport, OR (Figure 4). Naphthalene at Prineville, OR, Cobb, CA, and Columbia Gorge were also in volatilization Post-Wildfire (Figure S6). These results are discussed further in the “Post-Wildfire Comparative Analysis” section.

Figure 4.

Figure 4

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Individual p-PAH (top row) and a-PAH (bottom row) Post-Wildfire flux for each site. NOTES: Different scales between p-PAHs and a-PAHs. AQI data was not available for Newport, OR in 2018. Naphthalene and alkylated naphthalenes are in Figure S6.

A-PAHs

A-PAHs were also depositing or near equilibrium Post-Wildfire with the exception of retene, a proposed molecular marker of conifer wood combustion from wildfires, at Newport, OR (2018), Corvallis, OR (2018 and 2020), and Columbia Gorge (2018) (Figure 4).68 1-methylnaphthalene and 2-methylnaphthalene were also in volatilization at Prineville, OR (2018), Cobb, CA (2018), and the Columbia Gorge (2018 and 2020) (Figure S6). 2-methylphenanthrene and 2-methylanthracene were volatilizing at Newport, OR, but at a lower magnitude compared to the Wildfire smoke sample.

Post-Wildfire Comparative Analysis of PAH Soil–Air Exchange in Other Environments

Post-Wildfire diffusive flux values were also compared to soil-air diffusive flux of vapor-phase PAHs in other studies listed in Table S29. Similar to Wildfire flux, Post-Wildfire had the opposite effect on PAHs in flux. PAHs were generally in deposition, but at a lesser magnitude, while PAHs at locations in Bursa, Turkey, Kathmandu and Pokhara, Nepal, Shanghai, China, and the PRD in South China were in volatilization (Table S29).54−57 LMW PAHs and pyrene were mainly in flux Post-Wildfire compared to locations in Izmir, Turkey, Bursa, Turkey, the PRD in South China, and Shanghai, China that additionally had HMW PAHs in flux.29,54,56,57 These results demonstrate that wildfire events in the western U.S. can still have lasting effects on LMW PAH movement a month afterward that differs from residential, agricultural, industrial, and traffic sources globally.

Post-Wildfire Flux and AQI

To determine if the AQI of a wildfire associated with volatilization of PAHs from soil after a wildfire, we looked for a positive trend between the average AQI of a wildfire that occurred before the Post-Wildfire deployment for each site and the proportion of p-PAHs and a-PAHs in flux that were volatilizing after a wildfire. We did not see an association between AQI and either p-PAHs or a-PAHs in volatilization Post-Wildfire (r2 of 0.162 and 0.061) (Figure S7).

Post-Wildfire sampling took place between 17 and 34 days after a wildfire occurred. PAHs may have been volatilizing or degrading from the soil in the time when sampling did not take place, especially at Corvallis, OR and Cobb, CA where there were 25 or more days in between wildfire smoke occurrence and sampling. Table S31 shows estimated half-lives of PAHs in soils from the literature, which are highly variable for each PAH largely due to the temperatures and soil types assessed. Naphthalene, 1-methylnaphthalene, acenaphthene, and fluorene show some evidence of short half-lives (0.3–39 days) which may explain why we did not observe much volatilization after a wildfire.69−71 For example, Coover and Sims 1987 were not able to characterize half-lives for naphthalene and acenaphthylene due to a rapid rate lost in 60 days.71

Pre-Wildfire vs Post-Wildfire Flux

We questioned if there were long-term impacts from wildfire smoke on the soil-air exchange of PAHs. We evaluated the seven paired Pre- and Post-Wildfire samples to determine if there was a difference in diffusive flux of individual PAHs. Across all paired samples and sites, there were a total of 154 incidences where an individual p-PAH or a-PAH was in flux Pre-Wildfire and/or Post-Wildfire. 106 of these instances met one of the criteria (a PAH either having a 50% higher magnitude of flux between Pre-Wildfire and Post-Wildfire, a change in directionality, or only being detected Post-Wildfire) to suggest a difference in flux between events (Table S32–S33). The p-PAHs and a-PAHs that met these criteria are shown in Figure 5. A majority of naphthalenes (71.0%), p-PAHs (65.1%), and a-PAHs (72.4%) met at least one of the criteria (Table S31). At all locations except for Prineville, OR, p-PAHs and a-PAHs showed evidence of a difference between Pre-Wildfire and Post-Wildfire (Table S34 and Figure S8).

Figure 5.

Figure 5

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PAHs with a minimum of 50% difference in flux between Pre-Wildfire and Post-Wildfire for naphthalenes (top row), p-PAHs (middle row), and a-PAHs (bottom row) at each sample site ordered by increasing AQI. Red boxes around the different PAH types represent cases where a majority (≥50%) of the total PAHs in flux (naphthalenes, p-PAHs, or a-PAHs) have a minimum of 50% difference in flux. Sites are ordered from least to greatest AQI of the wildfire in between sampling events. NOTES: AQI data was not available for Newport, OR in 2018. Naphthalene and alkylated naphthalenes are separated due to larger order of magnitudes.

Generally, p-PAHs and a-PAHs displayed similar trends in how flux changed between Pre-Wildfire and Post-Wildfire events. For example, when p-PAHs at a site volatilized at a higher magnitude (or deposited less) Post-Wildfire compared to Pre-Wildfire, the a-PAHs tended to follow the same trend (Figure 5). However, naphthalenes tended to be more variable for how flux changed. Naphthalene, phenanthrene, acenaphthene, fluoranthene, 2-methylphenanthrene, 2-methylanthracene, and retene were volatilizing at greater magnitude nominally a month after a wildfire compared to before for at least one site. This suggests the potential for greater exposure to those PAHs following a wildfire from the air (Figure 5).

Limitations and Advantages

Wildfires are unpredictable regarding when and where they will occur, making it difficult to know exactly where to recruit, train, and send sampling materials to participants in advance. These challenges have resulted in small sample sizes, given considerations of participant availability and intersection with wildfire smoke exposure, as well as standard challenges around participant retention. However, an advantage of using PSDs is their ease of use and low maintenance for participants compared to active sampling. We were successfully able to increase the geographic extent of the study by sending PSD kits to participants to sample wildfire events across Washington, Oregon, and California, expanding over 966 km.

PSDs continuously sample, so they can capture episodic events like wildfires throughout the deployment. However, this can result in sampling during periods where wildfire smoke is no longer present after a wildfire. For example, each participant sampled for an average of 26 days, but we defined a sample as “Wildfire” if there were three consecutive days of heavy smoke. Thus, there may be Post-Wildfire influence in the Wildfire samples. It is important to keep in mind the safety of the participants throughout the Wildfire deployment. Air quality can remain poor, or nearby wildfires can result in evacuations, so it may not always be feasible for participants to collect PSDs immediately following a wildfire.

Other limitations to this study include the metadata used to define wildfire smoke events. AQI stations were not available at the same locations where sampling took place. Nearest AQI stations ranged from 11–97 km from the sampling locations. Air quality at stations farther away may risk not being exactly representative of air quality at the sampling location due to factors like prevailing wind directions or proximity to other sources of PM2.5. This may also influence why a trend between AQI of a wildfire and Post-Wildfire PAH volatilization was not observed. HMS covers a large geographic extent, but wildfires may be masked due to cloud cover or areas of a high sun angle.51 Despite their limitations, AQI and HMS has been used to define wildfire events in a previous study by Lee and Jaffe 2024.72

Only vapor-phase PAHs were analyzed in this study using lipophilic LDPE PSDs, which are optimal for compounds with a logKow between 4 and 10.73 There are opportunities for future studies to utilize different sampler types to investigate potentially toxic particulate-bound or more polar compounds in wildfire smoke. Understanding what other compounds are present in wildfire smoke and their fate in the environment could better inform future exposure assessments when people are exposed to these mixtures.

Acknowledgments

The authors want to thank Richard Scott, Kaley Adams, Lane Tidwell, Peter Hoffman, Clarisa Cabellero-Ignacio, Kaci Graber, and Caoilinn Haggerty for their help with gear acquisition and preparation, shipping, and laboratory processing; Michael Barton for data organization; and all members of the Food Safety and Environmental Stewardship Program at Oregon State University for identifying potential sampling locations. We especially wish to thank our participants in this study, without whom this work would not be possible.

Supporting Information Available

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acs.est.4c09139.

  • Additional details about sample preparation, collected metadata, analytical methodology, calculations, diffusive flux values, and data analysis (PDF)

This work was supported by the National Institute of Environmental Health Sciences (NIEHS) award numbers P42 ES016465, T32 ES007060, and P30 ES030287. This content is solely the responsibility of the authors and does not represent the official views of the NIEHS or NIH.

The authors declare no competing financial interest.

Special Issue

Published as part of Environmental Science & Technologyspecial issue “Wildland Fires: Emissions, Chemistry, Contamination, Climate, and Human Health”.

Supplementary Material

es4c09139_si_001.pdf (1.5MB, pdf)

References

  1. Dennison P. E.; Brewer S. C.; Arnold J. D.; Moritz M. A. Large wildfire trends in the western United States, 1984–2011. Geophys. Res. Lett. 2014, 41 (8), 2928–2933. 10.1002/2014GL059576. [DOI] [Google Scholar]
  2. Spreading like Wildfire – The Rising Threat of Extraordinary Landscape Fires; United Nations Environment Programme: Nairobi, 2022. [Google Scholar]
  3. Pausas J. G.; Keeley J. E. Wildfires and global change. Front Ecol Environ. 2021, 19 (7), 387–395. 10.1002/fee.2359. [DOI] [Google Scholar]
  4. Prichard S. J.; O'Neill S. M.; Eagle P.; Andreu A. G.; Drye B.; Dubowy J.; Urbanski S.; Strand T. M. Wildland fire emission factors in North America: synthesis of existing data, measurement needs and management applications. Int. J. Wildland Fire. 2020, 29 (2), 132–147. 10.1071/WF19066. [DOI] [Google Scholar]
  5. Ghetu C. C.; Rohlman D.; Smith B. W.; Scott R. P.; Adams K. A.; Hoffman P. D.; Anderson K. A. Wildfire Impact on Indoor and Outdoor PAH Air Quality. Environ. Sci. Technol. 2022, 56 (14), 10042–10052. 10.1021/acs.est.2c00619. [DOI] [PMC free article] [PubMed] [Google Scholar]
  6. Messier K. P.; Tidwell L. G.; Ghetu C. C.; Rohlman D.; Scott R. P.; Bramer L. M.; Dixon H. M.; Waters K. M.; Anderson K. A. Indoor versus Outdoor Air Quality during Wildfires. Environ. Sci. Tech Lett. 2019, 6 (12), 696–701. 10.1021/acs.estlett.9b00599. [DOI] [PMC free article] [PubMed] [Google Scholar]
  7. Rager J. E.; Clark J.; Eaves L. A.; Avula V.; Niehoff N. M.; Kim Y. H.; Jaspers I.; Gilmour M. I. Mixtures modeling identifies chemical inducers versus repressors of toxicity associated with wildfire smoke. Sci. Total Environ. 2021, 775, 145759 10.1016/j.scitotenv.2021.145759. [DOI] [PMC free article] [PubMed] [Google Scholar]
  8. Samburova V.; Connolly J.; Gyawali M.; Yatavelli R. L. N.; Watts A. C.; Chakrabarty R. K.; Zielinska B.; Moosmuller H.; Khlystov A. Polycyclic aromatic hydrocarbons in biomass-burning emissions and their contribution to light absorption and aerosol toxicity. Sci. Total Environ. 2016, 568, 391–401. 10.1016/j.scitotenv.2016.06.026. [DOI] [PubMed] [Google Scholar]
  9. Wegesser T. C.; Franzi L. M.; Mitloehner F. M.; Eiguren-Fernandez A.; Last J. A. Lung antioxidant and cytokine responses to coarse and fine particulate matter from the great California wildfires of 2008. Inhal Toxicol. 2010, 22 (7), 561–570. 10.3109/08958370903571849. [DOI] [PubMed] [Google Scholar]
  10. ATSDR: Toxicological Profile for Naphthalene, 1-Methylnaphthalene, and 2-Methylnaphthalene (Draft for Public Comment); Department of Health and Human Services, Public Health Service: Atlanta, GA: U.S., 2024; www.atsdr.cdc.gov/toxprofiles/tp67.pdf. [Google Scholar]
  11. Armstrong B.; Hutchinson E.; Unwin J.; Fletcher T. Lung cancer risk after exposure to polycyclic aromatic hydrocarbons: A review and meta-analysis. Environ. Health Persp. 2004, 112 (9), 970–978. 10.1289/ehp.6895. [DOI] [PMC free article] [PubMed] [Google Scholar]
  12. Brette F.; Shiels H. A.; Galli G. L. J.; Cros C.; Incardona J. P.; Scholz N. L.; Block B. A. A Novel Cardiotoxic Mechanism for a Pervasive Global Pollutant. Sci. Rep. 2017, 7, 41476. 10.1038/srep41476. [DOI] [PMC free article] [PubMed] [Google Scholar]
  13. Cakmak S.; Dales R. E.; Liu L.; Kauri L. M.; Lemieux C. L.; Hebbern C.; Zhu J. P. Residential exposure to volatile organic compounds and lung function: Results from a population-based cross-sectional survey. Environ. Pollut. 2014, 194, 145–151. 10.1016/j.envpol.2014.07.020. [DOI] [PubMed] [Google Scholar]
  14. Chepelev N. L.; Moffat I. D.; Bowers W. J.; Yauk C. L. Neurotoxicity may be an overlooked consequence of benzo[a]pyrene exposure that is relevant to human health risk assessment. Mutat Res-Rev. Mutat. 2015, 764, 64–89. 10.1016/j.mrrev.2015.03.001. [DOI] [PubMed] [Google Scholar]
  15. Huang X.; Xu X.; Dai Y.; Cheng Z.; Zheng X.; Huo X. Association of prenatal exposure to PAHs with anti-Mullerian hormone (AMH) levels and birth outcomes of newborns. Sci. Total Environ. 2020, 723, 138009 10.1016/j.scitotenv.2020.138009. [DOI] [PubMed] [Google Scholar]
  16. Liamin M.; Le Mentec H.; Evrard B.; Huc L.; Chalmel F.; Boutet-Robinet E.; Le Ferrec E.; Sparfel L. Genome-Wide Transcriptional and Functional Analysis of Human T Lymphocytes Treated with Benzo[a]pyrene. Int. J. Mol. Sci. 2018, 19 (11), 3626. 10.3390/ijms19113626. [DOI] [PMC free article] [PubMed] [Google Scholar]
  17. Sucker K.; Zschiesche W.; Aziz M.; Drews T.; Hummel T.; Raulf M.; Weiss T.; Bury D.; Breuer D.; Werner S.; Friedrich C.; Bunger J.; Pallapies D.; Bruning T. Naphthalene: irritative and inflammatory effects on the airways. Int. Arch Occ Env Hea. 2021, 94 (5), 889–899. 10.1007/s00420-020-01636-0. [DOI] [PMC free article] [PubMed] [Google Scholar]
  18. Sun K.; Song Y.; He F.; Jing M.; Tang J.; Liu R. A review of human and animal exposure to polycyclic aromatic hydrocarbons: Health risk and adverse effects, photo-induced toxicity and regulating effect of microplastics. Sci. Total Environ. 2021, 773, 145403 10.1016/j.scitotenv.2021.145403. [DOI] [PubMed] [Google Scholar]
  19. Zhang Y. Y.; Wang C. G.; Huang L. X.; Chen R.; Chen Y. X.; Zuo Z. H. Low-level pyrene exposure causes cardiac toxicity in zebrafish (Danio rerio) embryos. Aquat Toxicol. 2012, 114, 119–124. 10.1016/j.aquatox.2012.02.022. [DOI] [PubMed] [Google Scholar]
  20. Wassenaar P. N. H.; Verbruggen E. M. J. Persistence, bioaccumulation and toxicity-assessment of petroleum UVCBs: A case study on alkylated three-ring PAHs. Chemosphere. 2021, 276, 130113 10.1016/j.chemosphere.2021.130113. [DOI] [PubMed] [Google Scholar]
  21. Samburova V.; Zielinska B.; Khlystov A. Do 16 Polycyclic Aromatic Hydrocarbons Represent PAH Air Toxicity?. Toxics. 2017, 5 (3), 17. 10.3390/toxics5030017. [DOI] [PMC free article] [PubMed] [Google Scholar]
  22. Ramirez N.; Cuadras A.; Rovira E.; Marce R. M.; Borrull F. Risk Assessment Related to Atmospheric Polycyclic Aromatic Hydrocarbons in Gas and Particle Phases near Industrial Sites. Environ. Health Persp. 2011, 119 (8), 1110–1116. 10.1289/ehp.1002855. [DOI] [PMC free article] [PubMed] [Google Scholar]
  23. Cousins I. T.; Beck A. J.; Jones K. C. A review of the processes involved in the exchange of semi-volatile organic compounds (SVOC) across the air-soil interface. Sci. Total Environ. 1999, 228 (1), 5–24. 10.1016/S0048-9697(99)00015-7. [DOI] [Google Scholar]
  24. Fang Y.; Nie Z.; Die Q.; Tian Y.; Liu F.; He J.; Huang Q. Organochlorine pesticides in soil and air at and around a compound contaminated site: vertical distribution, soil–air exchange and risk evaluation. Stochastic Environmental Research and Risk Assessment. 2018, 32 (4), 1179–1188. 10.1007/s00477-017-1412-1. [DOI] [Google Scholar]
  25. Wang Y.; Li Z.; Tan F.; Xu Y.; Zhao H.; Chen J. Occurrence and air-soil exchange of organophosphate flame retardants in the air and soil of Dalian. China. Environ. Pollut. 2020, 265, 114850 10.1016/j.envpol.2020.114850. [DOI] [PubMed] [Google Scholar]
  26. Meharg A. A.; Wright J.; Dyke H.; Osborn D. Polycyclic aromatic hydrocarbon (PAH) dispersion and deposition to vegetation and soil following a large scale chemical fire. Environ. Pollut. 1998, 99 (1), 29–36. 10.1016/S0269-7491(97)00180-2. [DOI] [PubMed] [Google Scholar]
  27. Meharg A. A.; Shore R. F.; French M. C.; Osborn D. Dioxin and furan residues in wood mice (Apodemus sylvaticus) following a large scale polyvinyl chloride (PVC) fire. Environ. Pollut. 1997, 97 (3), 213–220. 10.1016/S0269-7491(97)00097-3. [DOI] [PubMed] [Google Scholar]
  28. Bozlaker A.; Muezzinoglu A.; Odabasi M. Atmospheric concentrations, dry deposition and air-soil exchange of polycyclic aromatic hydrocarbons (PAHs) in an industrial region in Turkey. J. Hazard Mater. 2008, 153 (3), 1093–1102. 10.1016/j.jhazmat.2007.09.064. [DOI] [PubMed] [Google Scholar]
  29. Demircioglu E.; Sofuoglu A.; Odabasi M. Particle-phase dry deposition and air-soil gas exchange of polycyclic aromatic hydrocarbons (PAHs) in Izmir. Turkey. J. Hazard Mater. 2011, 186 (1), 328–335. 10.1016/j.jhazmat.2010.11.005. [DOI] [PubMed] [Google Scholar]
  30. Donald C. E.; Anderson K. A. Assessing soil-air partitioning of PAHs and PCBs with a new fugacity passive sampler. Sci. Total Environ. 2017, 596, 293–302. 10.1016/j.scitotenv.2017.03.095. [DOI] [PMC free article] [PubMed] [Google Scholar]
  31. Dumanoglu Y.; Gaga E. O.; Gungormus E.; Sofuoglu S. C.; Odabasi M. Spatial and seasonal variations, sources, air-soil exchange, and carcinogenic risk assessment for PAHs and PCBs in air and soil of Kutahya, Turkey, the province of thermal power plants. Sci. Total Environ. 2017, 580, 920–935. 10.1016/j.scitotenv.2016.12.040. [DOI] [PubMed] [Google Scholar]
  32. Kaya E.; Dumanoglu Y.; Kara M.; Altiok H.; Bayram A.; Elbir T.; Odabasi M. Spatial and temporal variation and air-soil exchange of atmospheric PAHs and PCBs in an industrial region. Atmos Pollut Res. 2012, 3 (4), 435–449. 10.5094/APR.2012.050. [DOI] [Google Scholar]
  33. Liu G. Q.; Yu L. L.; Li J.; Liu X. A.; Zhang G. PAHs in soils and estimated air soil exchange in the Pearl River Delta, South China. Environ. Monit Assess. 2011, 173 (1–4), 861–870. 10.1007/s10661-010-1429-0. [DOI] [PubMed] [Google Scholar]
  34. Masih A.; Masih J.; Taneja A. Study of air-soil exchange of polycyclic aromatic hydrocarbons (PAHs) in the north-central part of India - a semi arid region. J. Environ. Monitor. 2012, 14 (1), 172–180. 10.1039/C1EM10567A. [DOI] [PubMed] [Google Scholar]
  35. Orazi M. M.; Arias A. H.; Oliva A. L.; Ronda A. C.; Marcovecchio J. E. Characterization of atmospheric and soil polycyclic aromatic hydrocarbons and evaluation of air-soil relationship in the Southwest of Buenos Aires province (Argentina). Chemosphere. 2020, 240, 124847 10.1016/j.chemosphere.2019.124847. [DOI] [PubMed] [Google Scholar]
  36. Pokhrel B.; Gong P.; Wang X. P.; Gao S. P.; Wang C. F.; Yao T. D. Sources and environmental processes of polycyclic aromatic hydrocarbons and mercury along a southern slope of the Central Himalayas. Nepal. Environ. Sci. Pollut R. 2016, 23 (14), 13843–13852. 10.1007/s11356-016-6443-5. [DOI] [PubMed] [Google Scholar]
  37. Wang C. F.; Wang X. P.; Gong P.; Yao T. D. Polycyclic aromatic hydrocarbons in surface soil across the Tibetan Plateau: Spatial distribution, source and air-soil exchange. Environ. Pollut. 2014, 184, 138–144. 10.1016/j.envpol.2013.08.029. [DOI] [PubMed] [Google Scholar]
  38. Rohlman D.; Samon S.; Allan S.; Barton M.; Dixon H.; Ghetu C.; Tidwell L.; Hoffman P.; Oluyomi A.; Symanski E.; Bondy M.; Anderson K. Designing Equitable, Transparent Community-Engaged Disaster Research. Citiz Sci. 2022, 7 (1), 36909292 10.5334/cstp.443. [DOI] [PMC free article] [PubMed] [Google Scholar]
  39. Anderson K. A.; Sethajintanin D.; Sower G.; Quarles L. Field trial and modeling of uptake rates of in situ lipid-free polyethylene membrane passive sampler. Environ. Sci. Technol. 2008, 42 (12), 4486–4493. 10.1021/es702657n. [DOI] [PubMed] [Google Scholar]
  40. Tidwell L. G.; Paulik L. B.; Anderson K. A. Air-water exchange of PAHs and OPAHs at a superfund mega-site. Sci. Total Environ. 2017, 603, 676–686. 10.1016/j.scitotenv.2017.01.185. [DOI] [PMC free article] [PubMed] [Google Scholar]
  41. Donald C. E.; Elie M. R.; Smith B. W.; Hoffman P. D.; Anderson K. A. Transport stability of pesticides and PAHs sequestered in polyethylene passive sampling devices. Environ. Sci. Pollut R. 2016, 23 (12), 12392–12399. 10.1007/s11356-016-6453-3. [DOI] [PMC free article] [PubMed] [Google Scholar]
  42. Donald C. E.; Scott R. P.; Wilson G.; Hoffman P. D.; Anderson K. A. Artificial turf: chemical flux and development of silicone wristband partitioning coefficients. Air Qual Atmos Hlth. 2019, 12 (5), 597–611. 10.1007/s11869-019-00680-1. [DOI] [PMC free article] [PubMed] [Google Scholar]
  43. Tidwell L. G.; Allan S. E.; O’Connell S. G.; Hobbie K. A.; Smith B. W.; Anderson K. A. PAH and OPAH Flux during the Deepwater Horizon Incident. Environ. Sci. Technol. 2016, 50 (14), 7489–7497. 10.1021/acs.est.6b02784. [DOI] [PMC free article] [PubMed] [Google Scholar]
  44. Anderson K. A.; Szelewski M. J.; Wilson G.; Quimby B. D.; Hoffman P. D. Modified ion source triple quadrupole mass spectrometer gas chromatograph for polycyclic aromatic hydrocarbon analyses. J. Chromatogr A 2015, 1419, 89–98. 10.1016/j.chroma.2015.09.054. [DOI] [PMC free article] [PubMed] [Google Scholar]
  45. Davie-Martin C. L.; Hageman K. J.; Chin Y. P. An Improved Screening Tool for Predicting Volatilization of Pesticides Applied to Soils. Environ. Sci. Technol. 2013, 47 (2), 868–876. 10.1021/es3020277. [DOI] [PubMed] [Google Scholar]
  46. Eek E.; Cornelissen G.; Breedveld G. D. Field Measurement of Diffusional Mass Transfer of HOCs at the Sediment-Water Interface. Environ. Sci. Technol. 2010, 44 (17), 6752–6759. 10.1021/es100818w. [DOI] [PubMed] [Google Scholar]
  47. Fernandez L. A.; Lao W. J.; Maruya K. A.; Burgess R. M. Calculating the Diffusive Flux of Persistent Organic Pollutants between Sediments and the Water Column on the Palos Verdes Shelf Superfund Site Using Polymeric Passive Samplers. Environ. Sci. Technol. 2014, 48 (7), 3925–3934. 10.1021/es404475c. [DOI] [PubMed] [Google Scholar]
  48. Gustafson K. E.; Dickhut R. M. Molecular Diffusivity of Polycyclic Aromatic-Hydrocarbons in Air. J. Chem. Eng. Data 1994, 39 (2), 286–289. 10.1021/je00014a020. [DOI] [Google Scholar]
  49. Schwarzenbach R. P.; Gschwend P. M.; Imboden D. M.. Environmental Organic Chemistry; John Wiley & Sons, 2003. [Google Scholar]
  50. Ferrari F.; Trevisan M.; Capri E. Predicting and measuring environmental concentration of pesticides in air after soil application. J. Environ. Qual. 2003, 32 (5), 1623–1633. 10.2134/jeq2003.1623. [DOI] [PubMed] [Google Scholar]
  51. McNamara D.; Stephens G.; Ruminski M.; Kasheta T.. In The Hazard Mapping System (HMS) - NOAA’S Multi-Sensor Fire and Smoke Detection Program Using Environmental Satellites. In 13th Conference on Satellite Meteorology and Oceanography, 2004.
  52. Gregoris E.; Argiriadis E.; Vecchiato M.; Zambon S.; De Pieri S.; Donateo A.; Contini D.; Piazza R.; Barbante C.; Gambaro A. Gas-particle distributions, sources and health effects of polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and polychlorinated naphthalenes (PCNs) in Venice aerosols. Sci. Total Environ. 2014, 476, 393–405. 10.1016/j.scitotenv.2014.01.036. [DOI] [PubMed] [Google Scholar]
  53. Li J.; Zhang G.; Li X. D.; Qi S. H.; Liu G. Q.; Peng X. Z. Source seasonality of polycyclic aromatic hydrocarbons (PAHs) in a subtropical city, Guangzhou, South China. Sci. Total Environ. 2006, 355 (1–3), 145–155. 10.1016/j.scitotenv.2005.02.042. [DOI] [PubMed] [Google Scholar]
  54. Liu Y.; Xie S. Y.; Zheng L. R.; Li T. T.; Sun Y. J.; Ma L. M.; Lin Z. F.; Grathwohl P.; Lohmann R. Air-soil diffusive exchange of PAHs in an urban park of Shanghai based on polyethylene passive sampling: Vertical distribution, vegetation influence and diffusive flux. Sci. Total Environ. 2019, 689, 734–742. 10.1016/j.scitotenv.2019.06.500. [DOI] [PubMed] [Google Scholar]
  55. Pokhrel B.; Gong P.; Wang X.; Chen M.; Wang C.; Gao S. Distribution, sources, and air-soil exchange of OCPs, PCBs and PAHs in urban soils of Nepal. Chemosphere. 2018, 200, 532–541. 10.1016/j.chemosphere.2018.01.119. [DOI] [PubMed] [Google Scholar]
  56. Sanli G.; Celik S.; Joubi V.; Tasdemir Y. Concentrations, phase exchanges and source apportionment of polycyclic aromatic hydrocarbons (PAHs) in Bursa-Turkey. Environ. Res. 2023, 232, 116344 10.1016/j.envres.2023.116344. [DOI] [PubMed] [Google Scholar]
  57. Wei Y. L.; Bao L. J.; Wu C. C.; He Z. C.; Zeng E. Y. Association of soil polycyclic aromatic hydrocarbon levels and anthropogenic impacts in a rapidly urbanizing region: Spatial distribution, soil-air exchange and ecological risk. Sci. Total Environ. 2014, 473, 676–684. 10.1016/j.scitotenv.2013.12.106. [DOI] [PubMed] [Google Scholar]
  58. Kim Y. S.; Lee M. J.; Seo D. S.; Kim T. H.; Kim M. H.; Lim C. H. Thirteen-week inhalation toxicity study of 1-methylnaphthalene in F344 rats. Toxicol Res-Ger. 2020, 36 (1), 13–20. 10.1007/s43188-019-00009-1. [DOI] [PMC free article] [PubMed] [Google Scholar]
  59. Swiercz R.; Wasowicz W.; Stetkiewicz J.; Gromadzinska J.; Majcherek W. 4-Week inhalation toxicity of 2-methylnaphthalene in experimental animals. Int. J. Occup Med. Env. 2011, 24 (4), 399–408. 10.2478/s13382-011-0036-9. [DOI] [PubMed] [Google Scholar]
  60. Donald C. E.; Nakken C. L.; Sorhus E.; Perrichon P.; Jorgensen K. B.; Bjelland H. K.; Stolen C.; Kancherla S.; Mayer P.; Meier S. Alkyl-phenanthrenes in early life stage fish: differential toxicity in Atlantic haddock (Melanogrammus aeglefinus) embryos. Environ. Sci-Proc. Imp. 2023, 25 (3), 594–608. 10.1039/D2EM00357K. [DOI] [PubMed] [Google Scholar]
  61. Peixoto M. S.; da Silva Junior F. C.; de Oliveira Galvão M. F.; Roubicek D. A.; de Oliveira Alves N.; Batistuzzo de Medeiros S. R. Oxidative stress, mutagenic effects, and cell death induced by retene. Chemosphere. 2019, 231, 518–527. 10.1016/j.chemosphere.2019.05.123. [DOI] [PubMed] [Google Scholar]
  62. Scott J. A.; Incardona J. P.; Pelkki K.; Shepardson S.; Hodson P. V. AhR2-mediated, CYP1A-independent cardiovascular toxicity in zebrafish (Danio rerio) embryos exposed to retene. Aquat Toxicol. 2011, 101 (1), 165–174. 10.1016/j.aquatox.2010.09.016. [DOI] [PubMed] [Google Scholar]
  63. Billiard S. M.; Querbach K.; Hodson P. V. Toxicity of retene to early life stages of two freshwater fish species. Environ. Toxicol. Chem. 1999, 18 (9), 2070–2077. 10.1002/etc.5620180927. [DOI] [Google Scholar]
  64. Wilson L. B.; McClure R. S.; Waters K. M.; Simonich M. T.; Tanguay R. L. Concentration-response gene expression analysis in zebrafish reveals phenotypically-anchored transcriptional responses to retene. Fron Toxicol. 2022, 4, 950503 10.3389/ftox.2022.950503. [DOI] [PMC free article] [PubMed] [Google Scholar]
  65. Magdoff F.; van Es H.. Building Soils for Better Crops Ecological Management for Healthy Soils; Department of Agriculture, 2021. [Google Scholar]
  66. Ukalska-Jaruga A.; Debaene G.; Smreczak B. O. D. Dissipation and sorption processes of polycyclic aromatic hydrocarbons (PAHs) to organic matter in soils amended by exogenous rich-carbon material. J. Soil Sediment. 2020, 20 (2), 836–849. 10.1007/s11368-019-02455-8. [DOI] [Google Scholar]
  67. Kim D. W.; Kim S. K.; Lee D. S. Relationship of pyrogenic polycyclic aromatic hydrocarbons contamination among environmental solid media. J. Environ. Monitor. 2009, 11 (6), 1244–1252. 10.1039/b900620f. [DOI] [PubMed] [Google Scholar]
  68. Ramdahl T. Retene - a Molecular Marker of Wood Combustion in Ambient Air. Nature. 1983, 306 (5943), 580–583. 10.1038/306580a0. [DOI] [Google Scholar]
  69. Sims R. C.; Overcash M. R. Fate of Polynuclear Aromatic-Compounds (Pnas) in Soil-Plant Systems. Residue Rev. 1983, 88, 1–68. 10.1007/978-1-4612-5569-7_1. [DOI] [Google Scholar]
  70. Park K. S.; Sims R. C.; Dupont R. R. Transformation of Pahs in Soil Systems. J. Environ. Eng-Asce. 1990, 116 (3), 632–640. 10.1061/(ASCE)0733-9372(1990)116:3(632). [DOI] [Google Scholar]
  71. Coover M. P.; Sims R. C. The Effect of Temperature on Polycyclic Aromatic Hydrocarbon Persistence in an Unacclimated Agricultural Soil. Hazard Waste Hazard. 1987, 4 (1), 69–82. 10.1089/hwm.1987.4.69. [DOI] [Google Scholar]
  72. Lee H.; Jaffe D. A. Impact of wildfire smoke on ozone concentrations using a Generalized Additive model in Salt Lake City, Utah, USA, 2006–2022. J. Air Waste Manage. 2024, 74 (2), 116–130. 10.1080/10962247.2023.2291197. [DOI] [PubMed] [Google Scholar]
  73. O’Connell S. G.; McCartney M. A.; Paulik L. B.; Allan S. E.; Tidwell L. G.; Wilson G.; Anderson K. A. Improvements in pollutant monitoring: Optimizing silicone for co-deployment with polyethylene passive sampling devices. Environ. Pollut. 2014, 193, 71–78. 10.1016/j.envpol.2014.06.019. [DOI] [PMC free article] [PubMed] [Google Scholar]

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