Mapping Indicators of Toxicity for Polycyclic Aromatic Compounds in

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Mapping Indicators of Toxicity for Polycyclic Aromatic Compounds in the Atmosphere of the Athabasca Oil Sands Region Narumol Jariyasopit, Tom Harner, Dongmei Wu, Andrew Williams, Sabina Halappanavar, and Ky Su Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b02058 • Publication Date (Web): 09 Sep 2016 Downloaded from http://pubs.acs.org on September 13, 2016

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Environmental Science & Technology

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Mapping Indicators of Toxicity for Polycyclic Aromatic Compounds in the Atmosphere of

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the Athabasca Oil Sands Region

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Narumol Jariyasopit1, Tom Harner1,*, Dongmei Wu2, Andrew Williams2, Sabina Halappanavar2,

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Ky Su1

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Air Quality Processes Research Section, Environment and Climate Change Canada, Toronto, Ontario, M3H 5T4 Canada 2

Environmental Health Science and Research Bureau, Health Canada, Ottawa, Ontario K1A 0K9, Canada Corresponding author: Dr. Tom Harner, phone: +1-416-739-4837, e-mail address: [email protected] For submission to Environmental Science & Technology

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ABTRACT Extracts of passive air samples collected from 15 passive sampling network sites across

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the Athabasca Oil Sands region were used to explore the application of in-vitro assays for

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mutagenicity (Salmonella mutation assays) and cytotoxicity (Lactate Dehydrogenase Assay) to

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assess the toxicity of the air mixture. The air monitoring of polycyclic aromatic compounds

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(PACs) and PAC transformation products, including nitrated polycyclic aromatic hydrocarbons

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(NPAHs) and oxygenated polycyclic aromatic hydrocarbons (OPAHs) was then linked to the

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potential toxicity of air. The PACs in air during April-May 2014 were elevated near mining

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activities and declined with distance from the source region, whereas NPAHs and OPAHs

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exhibited a more variable spatial distribution with highest levels in Fort McMurray. Overall, the

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air samples exhibited a weak mutagenicity. The highest indirect-acting mutagenicity was

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observed for sites closest to mining activities; however, the indirect-acting mutagenicity did not

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decline sharply with distance from mining areas. Indirect-acting mutagenicity was strongly 1 ACS Paragon Plus Environment


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correlated with levels of total PACs, benzo(a)pyrene equivalent mass, and OPAHs. Most of the

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samples exhibited cytotoxic potential but the magnitude of the response was variable across the

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sample region and did not correlate with levels of target analytes. This indicates that PACs and

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PAC derivatives were not a major contributor to the cytotoxicity observed in the air samples.

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Keywords: passive air sampling, in vitro assays, toxicity indicator map, mutagenicity,

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cytotoxicity, polycyclic aromatic compounds, PAC, PAH, PAH transformation products, PUF

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disk.

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INTRODUCTION

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Canada’s oil sands are the third largest petroleum reserves in the world. The main areas

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are located in northern Alberta, covering approximately 142,200 km2.1 It is estimated the

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Alberta’s reserves contain 170 billion barrels of recoverable oil.1 The Athabasca oil sands region,

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located in northeastern Alberta along the Athabasca river valley, is the largest of three oil sands

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deposits. Methods used to recover bitumen include open-pit surface mining and in-situ steam

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assisted extraction. The surface mining, accounting for 20% of the oil sands production, is

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limited to the Athabasca oil sands region where bitumen deposits are close to the surface.

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Impacts from oil sands development and operation, which have been rapidly expanding in the

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past decade, have been of interest from a variety of environmental aspects.

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Under the Joint Canada/Alberta Implementation Plan on Oil Sands Monitoring (JOSM),2

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passive air sampling monitoring network was established to measure PACs, alkylated PACs

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(Alk-PACs), dibenzothiophene (DBT), and alkylated DBTs.3 This passive air monitoring

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network is part of a broader ‘Deposition and Effects’ component; 3-5 which is part of an even

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more diverse and integrated monitoring effort that includes other environmental media including 2 ACS Paragon Plus Environment

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water, vegetation, soil, sediment and wildlife to better understand the cumulative environmental

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effects associated with mining activities.2 In the Athabasca oil sands region, the potential

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emission sources of these chemicals consist of open mining areas, upgrader facilities, mine

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fleets, tailing ponds, vehicular emissions, as well as forest fires. The petrogenic-derived PAC

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composition in environmental samples has a distinct feature in which Alk-PACs, DBT and Alk-

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DBTs dominate over unsubstituted polycyclic hydrocarbons (PAHs).6, 7 Therefore, PAH

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diagnostic ratios, for example the pyrogenic index, FLA/PYR, etc. are commonly used to

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distinguish environmental samples being contaminated by combustion vs petroleum sources.8-10

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However, it should be noted that deviation from the proposed values of the PAH diagnostic

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ratios as well as other uncertainties have been reported, and so they should be interpreted with

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caution.3 An earlier study demonstrated that both gas- and particulate-phase PACs and Alk-

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PACs were captured by the passive air samplers used in the network, at the same sampling rate

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of 5 m3/d.11 Since 2013, PAC derivatives, including nitrated polycyclic aromatic hydrocarbons

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(NPAHs) and oxygenated polycyclic aromatic hydrocarbons (OPAHs) have been included in the

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list of target analytes for the passive sampling program. The presence of PAC derivatives in air is

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due to secondary formation in the atmosphere and/or co-release with the parent PACs from

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incomplete combustion processes.12, 13 A number of NPAHs and OPAHs are listed as “probably’

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and “possibly” carcinogenic to humans 14, 15 Previous studies showed that they were significant

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contributors to the mutagenicity of air extracts.16-18 The importance of these PAC derivatives is

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also recognized because of their propensity to associate with particulate matter which enhances

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their deposition to terrestrial and aquatic surfaces and introduces them to the food chain.12

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Studies of environmental toxicity in the Athabasca oil sands region, have focused on contaminants present in oil sand process water 19-21 and river sediments22, 23 but toxicity

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associated with contaminants in air has been rarely investigated.24 Ambient air quality is often

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evaluated based on representative concentration limits of gas pollutants. For PAH mixtures in

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air, air quality objectives are based on benzo(a)pyrene (BaP), which is a known carcinogen.25

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However, these air guidelines do not include PAH transformation products, Alk-PACs, DBTs,

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and other toxic components of ambient air that may be elevated in air due to mining activities. In

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the current study the connection is made between chemical concentrations in air and the

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assessment of the corresponding toxicity. Toxicity indicator responses are assessed for the air

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mixture captured by polyurethane foam (PUF) disk type passive air samplers deployed at 15

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sampling sites in the Athabasca oil sands region during April to May 2014. This assessment of

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toxicity indicators takes into account the interactions among multiple chemical components,

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which consist of known and unknown chemicals, in the air mixture. The approach has been

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recently applied to explore the toxic potencies of air mixtures collected by PUFs in Eastern

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Europe.26 The objectives were to 1) measure PACs, NPAHs, and OPAHs in the oil sands

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samples, 2) assess toxicity indicator responses of the air extracts using two in-vitro bioassays

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including salmonella mutation assay and a mammalian lactate dehydrogenase cytotoxicity assay,

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and 3) develop toxicity indicator maps. Mapping of toxicity indicators is a pragmatic approach

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for assessing environmental health hazard as a complement to air concentration mapping.

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EXPERIMENTAL SECTION

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Chemicals and Materials. The monitored parent PACs and Alk-PACs were

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acenaphthylene (ACY), acenaphthene (ACE), fluorene (FLU), phenanthrene (PHE), anthracene

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(ANT), dibenzothiophene (DBT), retene (RET), fluoranthene (FLA), pyrene (PYR),

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benz(a)anthracene (BaA), chrysene (CHR), benzo(b)fluoranthene (BbF), benzo(k)fluoranthene

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(BkF), benzo(a)pyrene (BaP), perylene (PER), triphenylene (TRI), indeno(1,2,3-c,d)pyrene 4 ACS Paragon Plus Environment

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(IcdP), dibenzo(a,h)anthracene (DahA), benzo(ghi)perylene (BghiP), the C1-, C2-, C3-, and C4-

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alkylated PACs for NAP, FLU, PHE/ANT, FLA/PYR, BaA/CHR/TRI, and DBT. ΣPAHs is the

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sum of all individual unsubstituted PAH and DBT concentrations while ΣPACs is the sum of alk-

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PAC concentrations and the ΣPAHs. The 31 NPAHs and 8 OPAHs monitored are listed in Table

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SI.1.

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Deuterium-labeled PAHs, NPAHs, and 13C-phenanthrene were purchased from

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Cambridge Isotope Labs (Andover, MA), CDN Isotopes (Point-Claire, Quebec, Canada), and

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Chiron AS (Norway). The isotopically labeled recovery PAH and NPAH surrogates included

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2,6-dimethylnaphthalene-d12, acenaphthene -d10, 13C-phenanthrene, anthracene-d10,

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benzo(b)naphtho(2,1-d)-thiophene-d10, chrysene-d12, benzo(b)fluoranthene-d12, benzo(e)pyrene-

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d12, dibenz(a,h)anthracene-d14, indeno(123-cd)pyrene-d12, 1-nitronaphthalene-d7, 2-methyl-1-

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nitronaphthalene-d9, 5-nitroacenaphthene-d9, 9-nitroanthracene-d9, 3-nitrofluoranthene-d9, 1-

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nitropyrene-d9, and 6-nitrochrysene-d11. The labeled PAH and NPAH internal standards included

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fluorine-d10 and benz(a)anthracene -d12, 2-nitrobiphenyl-d9, and 2-nitrofluorene-d9. PUF disks

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(TE-1014, 14 cm diameter × 1.35 cm thick) were purchased from Tisch Environmental (Village

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of Cleves, OH).

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Sampling. The sampling sites and sample deployment have been described previously.3

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Briefly, PUF disks were pre-cleaned with accelerated solvent extraction (ASE) using acetone,

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petroleum ether, and acetonitrile before use. PUF disks were deployed from April-May 2014

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(~60 days) at 15 passive air sampling monitoring sites (Figure 1), which have been operated by

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the Wood Buffalo Environment Association (WBEA). A linear phase sampling rate of 5 m3/d

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was derived previously for both gas- and particle-phase PAHs.11 Two field blanks were included

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in the sampling. Duplicate samplers were deployed at sites L05 and L06. 5 ACS Paragon Plus Environment


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Sample preparation. PUF disks were cut in half. All the samples were extracted using

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the ASE with petroleum ether/acetone (75/25, v/v; 2 cycles). The half that was subjected to

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toxicity indicator testing was extracted without being spiked with recovery surrogates. The

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extracts were evaporated to dryness under a stream of N2 and the residue was dissolved in 500µl

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of dimethyl sulfoxide (DMSO). Because the extracts were not subjected to any purification or

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“clean-up” steps prior to conducting the toxicity assays, the responses reflect the whole air

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mixture which consists of multiple chemical components and their interactions. The second half

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of the PUF disk that was subjected to chemical analysis was spiked with recovery PAH and

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NPAH surrogates prior to the extraction. The extracts were purified using 5 g silica columns

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(Mega BE-SI, Agilent Technologies, New Castle, DE), eluted with petroleum ether/acetone

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(50/50, v/v) and were spiked with labeled PAH and NPAH internal standards. The analysis of

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PACs was conducted using a gas chromatography mass spectrometry (GCMS, Agilent 6890

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coupled with an Agilent 5975 MSD), using electron impact in selected ion monitoring (SIM)

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mode, on a DB-XLB column (30 m × 0.25 mm I.D., 0.25 µm film thickness, Agilent

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Technologies). The analysis of NPAHs and OPAHs was conducted using a GCMS (Agilent

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7890A coupled with and Agilent 7000 MSD), in electron capture negative ionization (ECNI), on

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a DB-5 column (30 m × 0.25 mm I.D., 0.25 µm film thickness, Agilent Technologies).

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Quality Assurance and Quality Control. All data were recovery and blank corrected.

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Average surrogate recovery for PACs, PAC derivatives were 86% (±22%) and 105% (±56%),

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respectively. The instrumental and method detection limits (IDL and MDL) are given in Table

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SI.1. Values below MDL were replaced by 2/3 of MDL.

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Toxicology Studies.

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Salmonella Mutation Assays. Salmonella strains TA98, Environmental Bio-detection 6 ACS Paragon Plus Environment

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Products Inc. lyophilized preparation, were used in the study. The full procedure for the

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Salmonella reverse mutation assays have been described in detail elsewhere by Maron and

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Ames.27 Briefly, 40 µl samples in DMSO, 2 ml melted top agar (45ºC), 0.5 ml of 0.1 M sodium

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phosphate buffer or rat liver S9 mix (an exogenous metabolic activation system based on rat liver

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enzymes, Moltox Inc. and EBPI), and 0.1 ml of bacteria were mixed and added to a plate

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containing Vogel-Bonner minimal agar. After the bacteria-containing agar was solidified for 30

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minutes, the plates were incubated at 37ºC in inverse position for 2 days. The bacterial revertant

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colonies (bacteria that have been mutated to grow on media deficient in histidine) were counted

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using a National Institute of Standard and Technology (NIST) Integrated Colony Enumerator. All

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air samples were tested in triplicate. The positive control concentrations for the assays with and

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without the rat S9 mix were 2 µg/plate of 2-aminoanthracene (2-AA) and 10µg/plate of 2-

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nitrofluorene (2-NF), respectively. Average revertant counts per plate of the positive controls

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were 2608 ± 1721 and 992 ± 181 for the assays with and without rat liver S9 mix, respectively.

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The negative control plates were plated with 40 µl of DMSO only. Average spontaneous

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revertant counts per plate of DMSO alone were 20 ± 6 and 15 ± 6 for the assays with and without

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rat liver S9 mix, respectively. Cell viability was verified by examining the background lawn at

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the highest sample concentration level and no adverse effect was observed (original bacterial

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colonies).

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Cell Culture and Treatment: Adenocarcinomic human alveolar basal epithelial cells

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(A549) were obtained from ATCC (Manassas, VA, USA). A549 cells are routinely used for

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characterizing pulmonary responses to environmental chemicals. Cells were cultured in F12K

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medium (ATCC, Manassas, VA, USA) supplemented with 10% (v/v) fetal bovine serum and 100

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µg/ml streptomycin, 100 U/ml penicillin (Invitrogen, Burlington, ON, Canada). Cells (4x104)



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were grown to ~90% confluence in 96 well plates at 37ºC in a humidified atmosphere of 5%

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CO2. Following a 24 h incubation period, the cells were treated with 1:125, 1:250, 1:500, and

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1:1000 dilutions of the air samples, each in triplicate. Controls were treated with DMSO (Fisher

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Scientific, New Jersey) only (n=3).

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LDH cytotoxicity assay: The LDH assay kit (Abcam, Toronto, ON, Canada) was used to

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measure the membrane integrity and release of cytoplasmic LDH into the medium. The assay

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was performed according to the manufacturer’s protocol. Cells treated with 10% cell lysis

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solution (all cells are killed resulting in maximum LDH release) were used as positive assay

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controls. Briefly, following 24 h exposure of cells to air sample extracts vehicle control (DMSO)

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or 10% cell lysis solution for 24 h, the culture plates were centrifuged at 600 x g for 10 min. The

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culture supernatants (10 µl) were transferred to wells of a new 96-well assay plate and mixed

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with 100 µl of LDH Reaction Mix. Samples were incubated for 30 min in the dark and the

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absorbance was measured at 450 nm. Changes in the absorbance reflect the degree of LDH

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release. Percent cellular toxicity was calculated as average increase in LDH released from cells

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treated with air extracts (LDH release is expected) from the matched controls treated with

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DMSO only (where LDH release is minimal or absent).

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RESULTS AND DISCUSSION

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Spatial Distribution of PACs. The passive sampling site map for this study is given in

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Figure 1. Site 11 (Lower camp) has been previously designated as the main oil sands source

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region given its proximity to oil sands related emissions.3 Individual PAC and PAC derivative air

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concentrations at all the sampling sites are given in Table SI.2 to Table SI.4 and expressed as

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mass collected by the passive samplers over the common deployment period. Due to its high

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volatility and abundance, naphthalene is not included in the reported ΣPACs values as well as the 8 ACS Paragon Plus Environment

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sum of the 15 US Environmental Protection Agency priority PAHs (ΣPAH15priority), which is

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consistent with our previous studies (Table 1).3, 11 Consistent with our early study, the highest

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ΣPACs mass (76,800 ng, which represents 99.7% of the total target compounds i.e.

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ΣPACs+NPAHs+OPAHs) was observed at L11 (Figure 1 and Table 1) that is centrally located

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near several mining-related sources and in an area of dense mining activity. Illustrated in Figure

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2, spatial distribution of ΣPACs showed a declining trend with increasing distance from L11

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(Table 1). On average, the Alk-PACs accounted for 94% of the ΣPACs for all the sampling sites,

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except for L06 (Residential site at Fort McMurray) and L14 (Anzac) where the Alk-PACs

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accounted for 85% and 59%, respectively (Figure SI.1). High PAHs/Alk-PACs ratios at the sites

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L06 and L14 were previously reported by Schuster et al.3 and Hsu et al.28 reported elevated

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PAHs in air at site L14. The elevated lower molecular weight PAH concentrations at L14 were

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attributed to the evaporation from nearby Gregoire lake.28 The pyrogenic index for the samples is

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given in Table 1. The pyrogenic index has been widely used in several studies to determine PAH

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compositional features.8, 29, 30 It is defined as the ratio of the sum of other three- to six-ring PAHs

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(ACY, ACE, ANT, FLA, PYR, BaA, BbF, BkF, BaP, PER, IcdP, DahA, BghiP, noting that

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biphenyl and benzo(e)pyrene were not monitored in this study) to the sum of the five target Alk-

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PACs (NAPs, PHEs, DBTs, FLUs, and CHRs). In bitumen and crude oils, unsubstituted PAHs

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are present in smaller amounts, relative to the alkylated homologues, therefore, a higher index

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value indicates a dominant contribution from pyrogenic sources (e.g. vehicle exhaust, wood

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burning) versus, for instance, petrogenic sources (e.g. bitumen, crude oils).8 The pyrogenic index

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of the oil sands samples and conventional crude oils were previously reported to range from

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0.01 to 0.037 while diesel soot samples ranged from 0.8 to 2.0.8 The pyrogenic index values at

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L06 (0.0706, town of Fort McMurray) and L14 (0.159, background site near Gregoire Lake)



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were relatively high compared to the other sampling sites where pyrogenic index values ranged

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from 0.01 to 0.04, indicating dominance of petrogenic sources (Table 1). The influence of

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pyrogenic (combustion) sources at site L06 (Fort McMurray) is further evident from the marker

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ratio FLA/PYR.9 The FLA/PYR ratio at L06 of 1.2 is substantially higher compared to a site in

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the source region, for instance site L11 (0.38, Lower Camp) (Table SI.2). Since FLA is enriched

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relative to PYR in combustion sources, larger values of the ratio indicate greater contribution

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from combustion. FLA/PYR values observed in urban areas were often closer to unity, indicating

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dominance of combustion sources of PACs.31-33 Furthermore, an analysis of raw Alberta oil

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sands samples revealed that the PYR concentration was 3-4 times higher than FLA,7 which

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would lead to a low value of the FLA/PYR ratio when petrogenic sources dominate.

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The detected NPAHs and OPAHs were 1-nitronaphthalene (1-NN), 1-methyl/2methyl-5-

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nitronaphthalene (1M5NN/2M5NN), 2-nitronaphthalene (2-NN), 9-nitroanthracene (9-NA), 2-

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nitrodibenzothiophene (2-NDBT), 2,8-dinitrodibenzothiophene (DNDBT), 9-fluorenone (FLO),

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9,10-anthraquinone (ANQ), 2-methyl-9,10-anthraquinone (MANQ), benzo(a)fluorenone (BaFL),

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benzo(b)fluorenone (BbFL), benzanthrone (BENZ), and benz(a)anthracene-7,12-dione (BaAD).

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NPAHs and OPAHs are of interest as some of them exhibit greater mutagenicity than the parent

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PAHs34, 35; previous studies have addressed concerns over toxic potential of air extracts

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containing atmospheric transformation products of PAHs.16, 36 However, this issue has so far not

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been addressed in the oil sands region and levels of PAH transformation products have so far not

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been reported for air. In the Athabasca oil sands region, the sum of ΣNPAHs and ΣOPAHs mass

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was highest at L06 (Fort McMurray) (Figure 1 and Table 1). Unlike the PAC spatial distribution

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which demonstrated a decline in the levels in air with increasing distance from L11, the patterns

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of ΣNPAHs and ΣOPAHs were variable across the sampling sites (Figure 2B).


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Thus far, there has been only one study investigating NPAH levels in air using passive air

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samplers.37 That study collected PUF disk type passive air samples (similar to those used in this

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study) at multiple locations in north China during July to October 2011 (90 days) and reported

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average NPAH masses of 341, 235, and 241 ng/sample for a megacity, town, and rural sites,

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respectively. These levels are one or two orders of magnitude higher than observed in the oil

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sands region (for 60 day deployments). Other studies have reported NPAH levels in air measured

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using active air samplers and reported in mass per unit of air sampling volume.18, 31, 38, 39 In order

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to put the NPAH and OPAH levels in oil sands region into a broader context, the ambient NPAH

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and OPAH concentrations were converted to more conventional air concentration units (volume

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based) using the sampling rate of 5 m3/d, the linear phase sampling rate that was derived

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previously for gas- and particle-phase PAHs.11 This results in an equivalent sample air volume of

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300 m3 for the entire sample. The resulting average concentrations in air for ΣNPAHs in the oil

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sands region was 20.9 pg/m3, ranging from below detection limit at R03 to 47.8 pg/m3 at L05

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(Mannix). These levels of ΣNPAHs are about an order of magnitude lower than the

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concentrations measured in urban areas, such as Los Angeles (851 pg/m3),31 Baltimore (318

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pg/m3),39 Marseille, France (710 pg/m3),38 and Beijing, China (792 pg/m3),18 but it is comparable

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to a rural area in Marseille, France (28 pg/m3).38

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Previous studies have shown that 2-NF and 2-NP, formed by gas-phase OH radical

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reactions, were the most abundant particle-associated NPAHs found in ambient air.31 For

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instance, Lin et al. reported an average of 2-NF value of 13.7 ng/sample for rural sites in north

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China, using passive samplers.37 However, in the current oil sands study 2-NF and 2-NP were

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below the detection limits in all samples. On the other hand, the compound 2-NDBT, which is

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rarely included or detected in other air monitoring programs, was detected with the highest level



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at L11 (6.9 ng/sample, Table SI.4). Overall, the NPAH concentrations in the oil sands region

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during April-May 2014 were generally lower than the values measured in the other urban areas.

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Although the parent PAH concentrations at sites that are closer to the source region (L11) are in

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the range of values reported at urban sites in European and North American countries,3 the

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relatively low abundance of the photochemically-derived NPAHs implies that atmospheric

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nitration potential is low at this oil sands impacted site.

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The OPAHs levels in air samples did not correspond to areas of intense mining activities

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(L11). The average value for ΣOPAHs across all sites was 129 ng/sample, ranging from 17.9

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ng/sample at R09 to 425 ng/sample at L06 with FLO being the dominant OPAH in all the

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samples. FLO is volatile and found primarily in the gas-phase.38 The town of Fort McMurray

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(L06) exhibited the highest FLO concentration of 206 ng/sample or approximately 644 pg/m3,

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using the same sample air volume conversion as for the NPAHs above. This result for FLO is

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higher than reported in other studies, for instance, Beijing (urban, 129 pg/m3),18 Paris (traffic

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site, 71 pg/m3),40 and Tokyo (urban, 21 pg/m3)32 but lower than observed in Marseille, France

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(urban, 3,577 pg/m3).38

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OPAHs are ubiquitous in urban ambient air and previous studies have associated them

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with fine particles released from combustion and products from reactions with atmospheric

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oxidants.12, 41, 42 Because the OPAHs are not associated with a specific pathway their relative

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concentrations in air cannot be used to differentiate sources. Nonetheless, the relatively high

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levels of OPAHs at L06 was consistent with its distinct PAH marker ratios (discussed earlier)

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which suggests that the L06 air sample was influenced by combustion sources (e.g. vehicular

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exhausts). These combustion sources could also be associated with primary emissions of the

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OPAHs. At the same time, OPAHs could also be formed as a transformation product in air due to 12 ACS Paragon Plus Environment

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local photochemical atmospheric aging that might be enhanced at L06 by gaseous oxidants (e.g.

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O3, NOx) which are related to the combustion sources. Yet a third possibility is that the OPAHs

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at site L06 are associated with regional transport and transformation (photochemical aging) of

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primary emissions from site 11 which is located upwind of L06.

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Mapping of Toxic Indicators. Air samples extracts were subjected to in vitro assays to

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assess potential toxicity endpoints including mutagenicity (both direct and indirect) and

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cytotoxicity. Approximately 4% of the sample was used for this purpose representing an

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equivalent sample air volume of 12 m3 air.

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Mutagenicity. Mapping of indirect-acting and direct-acting mutagenicity are shown in

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Figure 3 and Figure SI.2. Parent PACs as well as some PAH derivatives are known to be

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indirect-acting mutagens, requiring metabolic activation system, in the Salmonella assay (and in

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humans), to convert them into the active form.43, 44 On the other hand, some NPAHs and OPAHs

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exhibit direct-acting mutagenicity in the Salmonella assay without metabolic activation.45 The

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results of the mutagenicity tests are expressed as relative mutagenicity, which is the mean

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mutagenic response of a sample divided by the background response (assay blank tests) (Table 1)

292

in which all the samples were tested in triplicate. Mutagenic responses (number of revertants

293

(mutated bacteria colonies) per plate) are given in Table SI.6. In this case a relative mutagenicity

294

of 1.0 indicates the same response as for the assay blank. The mutagenicity of air sample field

295

blanks was close to unity for both assays (with and without metabolic activation) indicating that

296

the field blanks and sample matrix contained insignificant amount of mutagens. Overall, the

297

indirect-acting mutagenicity of the samples rarely exceeded two times of the background,

298

indicating a weak mutagenicity. Although the two-fold rule has been widely applied as an

299

indicator of a positive response, there have been published studies questioning validity of the 13 ACS Paragon Plus Environment


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rule.46-48 As such, the evidence calls for future study in which multiple doses are tested.

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However, results based on the single-dose testing in this study are informative in the aspect of

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spatial comparison of mutagenicity. The highest and statistically significant relative indirect-

303

acting mutagenicity of 1.8 was observed at site L11 corresponding to the site having the highest

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ΣPACs in air; however, the indirect-acting mutagenicity did not decline with increasing distance

305

from L11 in the same way that the PAC concentrations in air declined.

306

The PACs concentrations values in air that were used for determining correlations with

307

toxicity indicators were not blank corrected in order to reflect the actual amounts of chemicals in

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the samples which could potentially induce the toxic responses. PACs were shown to be a major

309

contributor to the indirect-acting mutagenicity of the samples given a strong and statistically

310

significant correlations between the indirect-acting mutagenicity and ΣPACs (R = 0.701, p-value

311

< 0.05) (Table SI.5). The correlation with the indirect-acting mutagenicity was slightly better for

312

the sum of unsubstituted PAHs (R = 0.758, p-value < 0.05) and was slightly lower for the ΣAlk-

313

PACs (R = 0.698, p-value < 0.05) (Table SI.5). To a lesser extent, the ΣOPAHs in air was

314

significantly correlated with the indirect-acting mutagenicity (R = 0.51, p-value < 0.05),

315

however, the ΣNPAHs was not significantly correlated with the indirect-acting mutagenicity.

316

Mutagenicity in the air samples can also arise from direct-acting mutagens, including

317

some NPAHs and OPAHs,12, 16, 49 that do not require metabolic activation (rat S9 mix) to be

318

mutagenic. Similar to the indirect-acting mutagenicity, L11 had the greatest direct-acting

319

mutagenic potential; however, all the samples generally induced low responses (Figure SI.2). It

320

is interesting that the L06 (Fort McMurray) sample that had the highest levels of sum of NPAH

321

and OPAH yielded considerably low direct-mutagenic response (i.e. relative mutagenicity of

322

1.2). The L06 sample was dominated by FLO and ANQ which accounted for 48% and 20%, 14 ACS Paragon Plus Environment

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respectively, of the sum of NPAHs and OPAHs. Despite the abundance of FLO in ambient air

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from studies discussed earlier, data on the mutagenicity of FLO are scarce. Although data on in-

325

vitro mutagenic potential of ANQ are available, the results are so far inconclusive due to issues

326

related to impurity in the tested standards.15 In this study the observation of low direct-mutagenic

327

response at L06 may suggest that the measured NPAHs and OPAHs and other unknown

328

chemicals are possibly weak direct-acting mutagens. The known potent direct-acting mutagens

329

among the NPAHs and OPAHs are 1-nitropyrene and dinitropyrenes, which were below the

330

detection limits for all samples in this study. In addition to the absence of potent direct-acting

331

mutagens, such low response could be attributed to suppression of the direct-acting mutagenicity

332

which may have been caused by the presence of PAHs. This was believed to promote the

333

interactions between aromatic compounds and suppress nitroreductase activity.17, 50 Although

334

direct-acting mutagenicity was also correlated with PACs (R = 0.600, p-value < 0.05), this

335

response is probably attributed to unknown chemicals that were co-emitted with or degraded

336

from the parent PACs, since PACs are indirect-acting mutagens that require exogeneous

337

bioactivation.

338

Cytotoxicity. Cytotoxicity is expressed as percentage increase in LDH release in cells

339

treated with air extracts relative to the control samples (DMSO). Field blank extracts did not

340

result in increased LDH release. Cells were exposed to different doses in triplicate and dose-

341

response curves for all the sites were calculated, which exhibited positive dose-response

342

relationships (Figure SI.3). For spatial comparison, the percent increases in LDH release

343

associated with the lowest dilution level (1:125) are shown in Figure 3B. The average percent

344

increase in LDH release (cytotoxicity) for samples was 120%, ranging from 3% at L13 to 223%

345

at R01 (Table 1). The increase for L11 and L06 (Fort McMurray) samples was 160% and 179%,



16

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respectively (Table 1). In general, most of the air samples exhibited cytotoxic potential, except

347

for two sites R05 and L13 in which the increase did not exceed 20% (Figure 3). Interestingly, the

348

samples that expressed higher cytotoxicity (>200%) were from remote sites including R08, R03,

349

and R01. The prevailing winds at R08 during the sampling period were from the oil sands area,

350

whereas they were typically from the southeast at site R03 and southwest and northeast at site

351

R01, representing remote forested areas.51 This result coupled with the lack of correlations

352

between the cytotoxicity and the PACs, NPAHs, and OPAHs (Table SI.5) concentrations

353

indicate that the PACs and PAH derivatives were not a major contributor to the cytotoxicity of

354

these samples. We also conducted a statistical analysis on the multiple-dose cytotoxicity results

355

to explore dose and site relationships among the sampling sites. The results are summarized in

356

Figures SI.4 (hierarchical cluster analysis represented as a tree diagram (Dendogram)) and SI.5

357

(a two-dimensional scaling plot). The dendogram reflects similarities and differences among

358

sites based on their dose-response curves (Figure SI.3). For instance the L06 duplicate samplers

359

and site R03 occur in the same “cluster” due to their very similar dose-response curves; whereas

360

the field blank is separate from all the other sites/clusters reflecting significantly different

361

cytotoxicity response. A similar clustering or grouping of sites is also seen in two-dimensional

362

scaling plot (SI.5). The results of the statistical analyses suggested that there were statistically

363

significant site-specific differences in the cytotoxic response induced by air extracts; however,

364

the responses did not correlate with the PACs, NPAHs or OPAHs content. Future studies are

365

needed to investigate contributions of other air contaminants (not measured in the current study)

366

to the cytotoxicity responses among sampling sites.

367 368

BaP equivalent (BaPeq) mass represents the contribution of carcinogenic PAHs in a mixture expressed in terms of an equivalent amount of BaP. In this study, individual BaP


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equivalent mass is determined by multiplying individual PAH concentration by its relative

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potency factor (RPF).52 Derivation of RPF values were mainly based on carcinogenicity bioassay

371

data. BaPeq of each sample is the sum of the individual BaPeq masses. RPF values for the PAHs

372

measured in this study are presented in Table SI.6. Despite the exclusion of Alk-PACs, the

373

spatial distribution of BaPeq was still consistent with that of the PACs showing the highest BaPeq

374

value at L11 (Table 1). BaPeq had stronger correlation with the indirect-acting mutagenicity (R =

375

0.72, p-value < 0.05) in comparison with the direct-acting mutagenicity (R = 0.59, p-value

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