Profiling quinones in ambient air samples collected from the Athabasca region (Canada)

Accepted Manuscript Profiling quinones in ambient air samples collected from the Athabasca region (Canada) Andrzej Wnorowski, Jean-Pierre Charland PII:

S0045-6535(17)31416-9

DOI:

10.1016/j.chemosphere.2017.09.003

Reference:

CHEM 19873

To appear in:

ECSN

Received Date: 6 June 2017 Revised Date:

31 August 2017

Accepted Date: 1 September 2017

Please cite this article as: Wnorowski, A., Charland, J.-P., Profiling quinones in ambient air samples collected from the Athabasca region (Canada), Chemosphere (2017), doi: 10.1016/ j.chemosphere.2017.09.003. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

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Profiling quinones in ambient air samples collected from the Athabasca

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region (Canada)

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Andrzej Wnorowski and Jean-Pierre Charland

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Environment and Climate Change Canada, Science and Technology Branch, Atmospheric

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Science and Technology Directorate, Air Quality Research Division, Analysis and Air Quality

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Section, , 335 River Rd., Ottawa, ON, K1V 1C7, Canada.

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Corresponding author: Tel.: 613-990-9613; fax: 613-990-8568; e-mail address:

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[email protected] (A. Wnorowski)

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Abstract

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This paper presents new findings on polycyclic aromatic hydrocarbon oxidation products–

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quinones that were collected in ambient air samples in the proximity of oil sands exploration.

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Quinones were characterized for their diurnal concentration variability, phase partitioning, and

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molecular size distribution. Gas-phase (GP) and particle-phase (PM) ambient air samples were

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collected separately in the summer; a lower quinone content was observed in the PM samples

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from continuous 24-h sampling than from combined 12-h sampling (day and night). The

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daytime/nocturnal samples demonstrated that nighttime conditions led to lower concentrations

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and some quinones not being detected. The highest quinone levels were associated with wind

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directions originating from oil sands exploration sites. The statistical correlation with primary

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pollutants directly emitted from oil sands industrial activities indicated that the bulk of the

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detected quinones did not originate directly from primary emission sources and that quinone

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formation paralleled a reduction in primary source NOx levels. This suggests a secondary

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chemical transformation of primary pollutants as the origin of the determined quinones.

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Measurements of 19 quinones included five that have not previously been reported in ambient air

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or in Standard Reference Material 1649a/1649b and seven that have not been previously

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measured in ambient air in the underivatized form. This is the first paper to report on quinone

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characterization in secondary organic aerosols originating from oil sands activities, to distinguish

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chrysenequinone and anthraquinone positional isomers in ambient air, and to report the

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requirement of daylight conditions for benzo[a]pyrenequinone and naphthacenequinone to be

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present in ambient air.

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Keywords: quinone OPAHs, secondary aerosol, gas–particle partition, distribution,

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daytime/nighttime variation.

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Highlights:

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-Nineteen quinones were determined without derivatization in ambient air.

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-Quinones were characterized by phase partitioning, size distribution, and diurnal content.

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-Night conditions led to lower concentrations and some quinones not being detected.

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-24-h vs. combined 12-h samplings resulted in the lower abundance of some quinones.

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-Measured quinones correlated strongly with oxidizing agents.

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1. Introduction

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The Athabasca oil sands region (AOSR) in Canada is the third-largest known reservoir of crude

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oil in the world and the largest of three major oil sands deposits in Canada. In 2016, it produced

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2.5 million barrels of bitumen per day via mining and in situ exploration (Alberta Energy Oil

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Sands, 2017). The increased petroleum exploration of the Athabasca oil sands since 1967 has

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raised health and environmental concerns over the release of and exposure to toxic compounds

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(Jautzy et al, 2013; Lundin et al., 2015; Zhang et al., 2016; Liggio et al., 2016; Ohiozebau et al.,

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2017). The primary sources of pollutant emissions associated with oil sands activities include

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open surface mines, bitumen extraction and upgrading, auxiliary transportation and traffic

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emissions, and the handling and storage of waste such as petroleum coke, tailing sands, and

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water. Primary pollutants which are directly emitted from various primary sources may be of

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health and environmental concern and hence are subject to routine monitoring. Such pollutants

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include volatile organic compounds (VOCs), CO, CO2, SO2, H2S, NOx, NH3, CH4, O3,

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hydrocarbons, particulate matter and metals (NAPS, 1969; WBEA, 1996; AEP, 1997). Oil sands

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exploration has also been recognized as a concerning source of polycyclic aromatic hydrocarbon

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(PAHs) emissions (Kelly et al., 2009; Kurek et al., 2013; Hsu et al., 2015) and has prompted

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continued monitoring and numerous systematic research studies and programs to assure

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environmentally responsible development of the oil sands (HEAP, 2000; HEMP, 2007; RAMP,

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2008; Yang et al., 2011; JOSM, 2012; Percy et al., 2012; NAPS, 1969; Watson et al., 2013;

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Schuster et al., 2015; Zhang et al., 2015; Bari and Kindzierski, 2015). However, primary

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quinone emissions and their secondary formation from the oxidation of PAHs and partitioning

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into secondary organic aerosols (SOA) have not been studied.

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Quinones are fully conjugated cyclic dione structures derived from aromatic compounds via the

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conversion of an even number of –CH= groups into –C(=O)– groups with the rearrangement of

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double bonds as necessary (IUPAC, 1997). While a variety of quinones are naturally occurring

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(Rogge et al., 1993), most of those present in the environment arise from anthropogenic primary

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sources (mainly the combustion of fossil fuels) and secondary reactions of PAH precursors

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(Lundstedt et al., 2007; Layshock et al., 2010; Keyte et al., 2013). While certain quinones are

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used as dyes and reagents or for healing (Newman and Cragg, 2012; López et al., 2015), other

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quinones may be toxic or carcinogenic to humans (Shang et al., 2014, Chen et al., 2015) or

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harmful to the environment (Charlson et al., 1992; Tiea and Cao, 2009). Toxicological studies

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have reported quinones to be more toxic than their parent PAHs because they exhibit direct

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mutagenic potency without enzymatic activation (Durant et al., 1998; Lundstedt et al., 2007;

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Wang et al., 2011). Quinones are reactive in the atmosphere and living organisms owing to their

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electrophilic properties and redox cycling abilities (Bolton et al., 2000; López et al., 2015). Solar

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photo-activation and chemical oxidation of PAHs can also lead to the generation of oxy-PAHs,

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including quinones (Lampi et al., 2006; Keyte et al., 2013 and references therein), in secondary

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organic aerosol at concentrations that are often higher than those of the unreacted parent

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compounds (Layshock et al., 2010; Bandowe et al., 2014). This can significantly contribute to

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the increased toxicity of ambient air containing high levels of PAHs (Lemieux et al., 2007). In

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addition, some oxy-PAHs, including quinones, are known to be a byproduct of bioremediation at

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PAH-contaminated sites (Lundstedt et al., 2003, 2007). Therefore, it would seem advantageous

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to include quinone monitoring at industrial and PAH-contaminated sites.

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A search of the environmental literature indicates that quinones are one of the least studied

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compounds. This is primarily due to the challenges in analytical methods (Lintelmann et al.,

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2006; Alves et al., 2017) and the lack of regulatory requirements, and consequently, systematic

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studies. Thus, there is a paucity of information concerning the presence, fate, and effect of

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quinones in ambient air (Wilson et al., 1995). Although the literature is sparse, sufficient Page 3 of 40

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evidence of their presence has accumulated in recent years to justify greater environmental

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attention to quinones (Lundstedt et al., 2007; Layshock et al., 2010; Bandowe et al., 2014; Shang

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et al., 2014, Walgraeve et al., 2015).

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Although parent PAHs have been routinely monitored in ambient air (US EPA, 1999; CSTEE,

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2001; Wnorowski et al., 2006; Lammel, 2015), only a few studies have reported ambient quinone

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concentrations (Walgraeve et al., 2010; Delgado-Saborit et al., 2013; Nocun and Schantz, 2013;

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Alam et al., 2014; Walgraeve et al., 2015). While PAHs are commonly analyzed by gas

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chromatography/mass spectrometry (GC-MS), using this method for quinones in the same

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samples presents a difficult analytical challenge (Lintelmann et al., 2005; Wingfors et al., 2011).

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In order to improve the analytical sensitivity for quinone detection, some studies have used

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derivatization techniques (Cho et al., 2004; Jakober et al., 2007). However, not all quinones are

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derivatized efficiently, and thermally labile quinones can be transformed and/or degraded during

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derivatization and analysis (Chung et al., 2006). HPLC-MS/MS offers many advantages over

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GC-MS in quinone analysis. It is an alternative method that does not require derivatization of

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compounds that are thermally labile, nonvolatile or have highly polar functional groups, thus

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resulting in less labor and cost intensive maintenance. In addition, HPLC allows for

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simultaneous analysis of various polarity analytes and for simplified sample cleanup (Cai et al.,

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2009; Walgraeve et al., 2010).

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A high-pressure liquid chromatography–tandem mass spectrometry (HPLC-MS/MS)-based

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method (Wnorowski and Charland, 2013) is presented in the current study to quantify a suite of

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one- to five-ring quinones in ambient air gas and particle phases downwind of Fort McKay South

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(Canada) near oil sands mining and processing sites. This improved method was used to

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measure the quinone levels and provide insight into the primary emission sources and

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contribution to secondary organic aerosols. The levels of phase-separated quinones from

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ambient air samples systematically monitored for 24- and 12-h periods have not previously been

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reported in the scientific literature.

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The objectives of the present study were to evaluate the occurrence, phase partitioning,

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molecular size distribution, and variation patterns of 19 quinones in daytime, nighttime, and Page 4 of 40

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daily ambient air samples in order to elucidate their origins and sources in the proximity of oil

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sands and to better understand their behavior and contribution to organic aerosols.

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2. Experimental section

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The following chemicals were investigated in this study (compounds properties listed in Table

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S1, Supplementary Information (SI)): 1,4-benzoquinone (1,4-BQ); 1,2-naphthoquinone (1,2-

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NAPQ); 1,4-naphthoquinone (1,4-NAPQ); 2-methyl-1,4-naphthoquinone (2-M14NAPQ); 2-

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hydroxy-1,4-naphthoquinone (2-H14NAPQ); 5-hydroxy-1,4-naphthoquinone (5-H14NAPQ);

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1,2-acenaphthenequinone (1,2-ACNAQ); 1,4-anthraquinone (1,4-ANTQ); 9,10-anthraquinone

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(9,10-ANTQ); 9,10-phenanthraquinone (9,10-PHEQ); 2-methylanthraquinone (2-MANTQ); 1,2-

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aceanthrenequinone (1,2-ACANQ); 7,12-benz[a]anthraquinone (7,12-BANTQ); 1,4-

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chrysenequinone (1,4-CHRYQ); 5,6-chrysenequinone (5,6-CHRYQ); 5,12-naphthacenequinone

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(5,12-NAPACQ); 1,6-benzo[a]pyrenedione (1,6-BaPyQ); 4,5-benzo[a]pyrenedione (4,5-

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BaPyQ); and 6,12-benzo[a]pyrenedione (6,12-BaPyQ). Quinone standards with a purity greater

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than 94.5% were purchased from 3B Pharmachem (Wuhan, China), CHEMOS GmbH & Co. KG

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(Regenstauf, Germany), Sigma Aldrich (Oakville, ON), and MRI (Kansas City, MO).

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Deuterated surrogates were purchased from CDN Isotopes (Point-Claire, Québec) and CIL

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(Tewksbury, MA). Additional information on standards preparation, calibration, detection

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limits, and quality control are provided in the SI.

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2.1. Sampling information

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Ambient air was sampled as part of the National Air Pollution Surveillance Program (NAPS,

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1969) in Canada. Two simultaneously operated high-volume (HiVol) samplers collected the

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particle-phase (PM30) (on a Teflon-coated borosilicate glass fiber filter) and gas-phase (on two

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polyurethane foam cartridges (PUFs)) associated compounds for 24-h or 12-h depending on the

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sampling schedule (Table S2) to investigate daytime, nocturnal, and daily concentrations. The

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sampling location (AMS13 Fort McKay South, Alberta, Canada) was approximately 15–20 km

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north and south from current oil sands-related industrial activities and away from any local direct

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emission sources (Figure S1). After collection, the PUFs and filter were secured, shipped, and

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stored until analysis at 0 °C for no more than 24-h before extraction.

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Evidence of a breakthrough in PUF sorbents resulting in an underestimation of gas-phase

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concentrations for compounds with a vapor pressures greater than phenanthrene has previously

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been reported (Possanzini et al., 2004; Su et al., 2006). Consequently, quantitative data derived

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for these compounds (with vapor pressures greater than phenanthrene) from the gas phase were

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not included in the present study (values for 1,4-BQ, 1,2-NAPQ, 1,4-NAPQ, and 2-M14NAPQ

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enclosed in brackets in Tables 1-3 are only used to indicate trends). No further adjustment nor

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sampling correction was applied to the obtained data. The sampling details are further presented

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in the SI.

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2.3. Sample preparation and analysis

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Prior to extraction, filters and PUFs (field samples, field blanks, and laboratory controls) were

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spiked with 100 µL of recovery standards (10 ng/µL). Each filter and PUF-pair were Soxhlet

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extracted for 24-h. Each extract was fractionated using silica gel column chromatography: first

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with hexane (F1-aliphatic sample), then with benzene (F2-PAH sample), and finally with

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methanol (F3-MeOH sample). Quinones were determined from the F3-MeOH fractions in

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MS/MS mode.

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An Agilent 1200SL high-performance liquid chromatograph was coupled to a ThermoFisher

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TSQ Quantum Discovery Max Triple-Quad MSD for the separation and detection of analytes of

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interest. Chromatographic separation was achieved with a Phenomenex HALO Phenyl-Hexyl

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150 × 3.0 mm × 2.7 µm column kept at 50 °C. A 2 µL sample injected volume was analyzed

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using a four-solvent gradient for isomer baseline separation, as described elsewhere (Wnorowski

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and Charland, 2013). Mobile phase solvents (water, methanol, 2-propanol, and acetonitrile) with

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Accusolv grade purity were purchased from Anachemia (Montréal, Québec). Argon of Grade

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5.0 for the collision cell was purchased from Praxair (Ottawa, Ontario). A Parker-Balston

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Nitrogen Generator N2-14A (Cleveland, OH) was used to produce nitrogen (99.5%) for the

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MSD sheath and auxiliary gases. Data were acquired under negative-mode atmospheric pressure

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chemical ionization (APCI) in selected reaction mode with Q1 = 0.2 and Q3 = 0.7 FWHM for

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enhanced resolution and selectivity. The Q2 CID was set to 1.7 mTorr argon. For each

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compound, the tube lens offset and collision energy were optimized. All studied analytes

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showed pseudo-molecular ion and two fragment ions under negative-mode APCI. The quality

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criteria for compound confirmation by MS/MS were based on identifying the parent and two

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product ions. The transition fragment with the higher intensity was used as a quantifier, and the

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second fragment was used as a qualifier. Typical LC-MS/MS chromatograms of the quinone

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standards and ambient air sample content are presented in Figure S2.

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2.4. Method validation and uncertainties

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Uncertainties as precision were calculated from the %RSD (relative standard deviation). The

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repeatability of the values for the ambient air samples was used to derive the uncertainties. The

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reproducibility %RSD from seven consecutive injection repetitions ranged from 7% for 1,4-BQ

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to 13% for 7,12-BANTQ. NIST SRM1649b reported concentrations of 1.8 mg/kg for 1,4-ANTQ

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and 3.6 mg/kg for 7,12-BANTQ, correlated well with the experimental values in the present

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study of 1.61±0.12 mg/kg for 1,4-ANTQ and 4.07±0.61 mg/kg for 7,12-BANTQ. This reflects

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an accuracy error of less than 14% (Table S3). The data in Table S3 also indicate good

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agreement between the experimental values and quinone content in NIST SRM 1649b of the

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literature. The 19 studied quinones had recoveries between 65% (7,12-BANTQ) and 88%

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(ANTQ isomers) (Table S4).

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Additional details regarding the sample preparation and chromatographic and spectroscopic

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analyses can be found in the SI.

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2.5. Environmental and atmospheric conditions during sampling

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Meteorological data (temperature (TEMP), relative humidity (RH), wind speed (WSP)), wind

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direction (WD), and primary pollutant concentrations (PM2.5, SO2, NO2, NO, O3, total

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hydrocarbons content (THC), total suspended particulates (TSP)) were recorded during sampling

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(Table S5 and Figure S3) and used for the correlation study (Table S6).

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3. Results and discussion

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3.1. Characterization of quinone content in relation to the sampling duration

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3.1.1. Quinone content resulting from 24-h sampling. The comparison of individual levels of

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quinones and their corresponding parent PAHs have already been reported elsewhere

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(Wnorowski, A., 2017) therefore, the current study is focusing on total quinone levels in diurnal

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and 24-h samples in the GP and PM.

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The summed daily quinone concentrations are graphically presented in Figure 1A and listed

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quinones, 2-H14NAPQ, 4,5-BaPyQ, and 6,12-BaPyQ were not detected in either phase, and

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5,12-NAPACQ and 1,2-ACNAQ were below their quantified limits in the GP. In addition, a

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considerable fraction of quinones was detected in the PM (total concentration ratio of 4:1 =

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PM:GP). Although quinones were present in both phases, those in the GP may have been subject

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to further chemical transformations at an accelerated rate relative to those associated with

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particles (Vione et al., 2004; Goldfarb and Suuberg, 2008; Lammel et al., 2009). This seems

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consistent with their higher concentrations observed in the studied PM sample.

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Table 1: Gas- and particle-phase ambient quinone content measured in 24-h time integrated

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samples (n = 19 days)

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Gas Phase [ng/m3]

Particle Phase [ng/m3]

Max

Mean

(0.180)a

(10.342)

(2.457)

1,2-NAPQ

(1.692)

(6.402)

1,4-NAPQ

(0.006)

2-M14NAPQ

Min

Max

Mean

Median

(1,769)

(22.114)

0.646

32.456

4.931

1.954

73.962

NAd

(3.556)

(2.574)

(10.667)

1.109

5.880

2.197

1.494

13.182

NA

(0.215)

(0.059)

(0.032)

(0.353)

0.086

0.247

0.166

0.166

0.333

NA

(0.012)

(0.052)

(0.017)

(0.012)

(0.206)

0.022

0.161

0.075

0.073

1.199

NA

2-H14NAPQ

-

-

-

-

NDc

-

-

-

-

ND

NA

5-H14NAPQ

0.160

1.074

0.587

0.528

1.761

0.357

2.782

1.200

0.461

3.600

0.5

1,2-ACANQ

0.016

0.075

0.050

0.055

0.201

0.053

2.214

0.680

0.227

2.722

0.1

1,4-ANTQ

0.018

0.187

0.065

0.055

0.518

0.065

0.419

0.175

0.172

1.747

0.3

9,10-ANTQ

0.008

0.140

0.027

0.015

0.296

0.009

0.481

0.117

0.075

1.874

0.2

9,10-PHEQ

0.009

0.091

0.049

0.048

0.196

0.023

0.033

0.028

0.028

0.056

3.5

2-MANTQ

0.014

0.016

0.015

0.014

0.044

0.010

0.078

0.021

0.011

0.220

0.2

EP -

-

-

b

0.124

0.202

0.171

0.187

0.513

NA

7,12-BANTQ

0.015

0.137

0.028

0.022

0.439

0.015

0.853

0.140

0.024

2.668

0.2

1,4-CHRYQ

0.023

0.030

0.027

0.027

0.081

0.021

0.083

0.053

0.055

0.159

0.5

5,6-CHRYQ

0.111

0.152

0.134

0.138

0.401

0.029

0.152

0.101

0.114

0.913

0.4

b

-

-

-

b

0.155

0.185

0.172

0.175

0.515

NA

1,6-BaPyQ

0.014

0.018

0.016

0.016

0.048

0.016

0.260

0.179

0.260

0.536

0.1

4,5-BaPyQ

-

-

-

-

ND

-

-

-

-

ND

NA

6,12-BaPyQ (Total w. Excluded) Total w/o Excluded a

-

-

-

-

ND

-

-

-

-

ND

NA

5,12-NAPACQ

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Total

b

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1,2-ACNAQ

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Gas / Particle Partition

Total

1,4-BQ

Median

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Min

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Quinone

(37.325)

(104.200)

3.985

15.524

Values enclosed in brackets were likely underestimated by some unknown factor because of

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b

S/N < 10; could not be quantified.

c

ND: No analytical signal detected.

d

NA: could not be calculated.

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Figure 1: Comparison of daily trends between the (A) daily total concentration of 19-quinones

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(purple dash line: quinone median concentration 0.45 ng/m3), (B) daily total concentration of 7-

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parent PAHs (blue dash line: PAH median concentration 1.64 ng/m3) (source: Wnorowski,

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2017), (C) daily NO2 concentration, for the 24-h composite (GP+PM) samples (n = 19 days) in

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relation to ambient air temperature.

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Table 1 also indicates that the quinone content was dominated by lighter compounds, such as

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1,4-BQ, 1,2-NAPQ, 5-H14NAPQ, and 1,2-ACANQ. In general, the quinone concentrations

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determined in both phases increased with decreasing molecular weight.

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To the best of our knowledge, no investigation has been carried out on quinone levels resulting

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from oil sands exploration. When the present study results were put in a global context with

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quinone measurements from various regions (Table S7), it became evident that quinones not only

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occur downwind of oil sands activities but also can be detected in relation to other petrogenic

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and industrial sources in various seasons and locations. With the exception of 1,4-BQ and 1,2-

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NAPQ, the determined average quinone concentrations in the study were in good agreement with

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ranges reported in the literature (Table S7) for samples collected in the summer months: traffic

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roadside in Munich (Lintelmann et al., 2005), the Athens city center (Valavanidis et al., 2006), a

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residential site in Fresno (Chung et al., 2006), various sites in the Chamonix and Maurienne

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valleys (Albinet et al., 2008a), a traffic site in London (Delgado-Saborit et al., 2013), various

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sites in Chiang Mai (Walgraeve et al., 2015) and coastal cities in Saudi Arabia (Harrison et al.,

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2016). While they had an insignificant contribution to the summed concentrations of

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determined quinones, the 7,12-BANTQ concentrations associated with the PM reported in the

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current study were on average four times higher than at suburban sites and 1.5 times higher than

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at traffic sites reported by Ringuet et al. (2012) for Palaiseau and Paris, Schnelle-Kreis et al.

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(2007) for urban Augsburg, and Walgraeve et al. (2015) for urban Chiang Mai. Other studies for

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traffic and urban sites in Grenoble (Tomaz et al., 2017), in three Southern European cities (Alvez

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et al., 2017) and in Rome (Di Filippo et al., 2015) reported similar or higher concentrations of

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7,12-BANTQ. However, comparing determined quinone concentrations with published reports

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is not straightforward. As noted by Delgado-Saborit et al. (2013) most investigators only report

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the PM content without isomer differentiation.

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The most abundant of the 19 quinones from the 24-h sampling in the vicinity of oil sands

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exploration were 1,4-BQ and 1,2-NAPQ. Although their GP levels could not be quantified

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accurately because of a possible breakthrough during the gas-phase sampling, the PM contents

275

significantly exceeded those reported for the London traffic roadside and other sites (Delgado-

276

Saborit et al., 2013). This distinct fingerprint of 19 quinones sets a reference point for all future Page 12 of 40

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277

quinone monitoring studies and evaluations of emissions originating from oil sands exploration,

278

bitumen, and possibly petroleum processing.

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Figure 2: Variations in daytime/nocturnal quinone concentrations in the (A) PM and (B) GP (n =

283

18 days).

284

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285

3.1.2. Quinone content resulting from 12-h sampling. In addition to 24-h sampling, the

286

daytime and nocturnal time series of quinone concentrations in the GP and PM were

287

concurrently examined. Figures 2A and 2B and Tables 2 and 3 present the daytime and

288

nocturnal quinone concentrations. The summed concentrations derived from this dataset

289

indicate that quinones determined in the daytime from the GP (Table 2) were approximately 30% Page 14 of 40

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more abundant than those from corresponding samplings at night (Table 3), and were 40% more

291

abundant when collected from the PM in the daytime (Table 2) than at night (Table 3). The

292

higher daytime concentrations may result from number of factors promoting quinone formation,

293

such as the release of primary pollutants associated with daytime-related industrial activities,

294

oxidative transformation of PAHs, higher temperatures (on average, +5 °C during the day than at

295

night) affecting the kinetics, higher concentrations of ozone (on average, 1.6 times than recorded

296

at night) and OH radicals (Keyte et al., 2013 and therein). However, data in Figure 2A also

297

indicate significantly higher nocturnal than daytime quinone levels for August 16, 17, and 18.

298

The reason for the higher nocturnal quinone content for these days is unclear, as oxidant levels

299

(NOx and O3) seemed to be within average concentration measured for this study (Figure S3 and

300

Table S5). It may reflect a low supply of PAHs for quinone generation, as parent PAH content

301

in GP for the corresponding days was also low (Figure S4B), reflecting possibly a higher

302

dilution/dispersion effect due to stronger wind speeds observed on these dates compared to the

303

other sampling days (Figure S3).

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304 305

Table 2: Daytime ambient quinones measured in 12-h time-integrated samples (n=17 days) Gas-Phase (ng/m3)

Particle-Phase (ng/m3)

Mean

Median

Total

Min

Max

Mean

Median

Total

1,4-BQ

a

(0.183)

(5.653)

(1.682)

(0.435)

(13.457)

0.419

5.814

2.639

2.443

39.579

NAc

1,2-NAPQ

(0.312)

(3.663)

(1.491)

(0.996)

(5.966)

0.259

3.720

1.457

0.925

5.828

NA

1,4-NAPQ

(0.134)

(0.134)

(0.134)

(0.134)

(0.134)

0.083

3.942

0.769

0.187

5.385

NA

2-M14NAPQ

(0.012)

(4.526)

(0.450)

(0.072)

(6.296)

0.013

0.320

0.100

0.057

1.396

NA

2-H14NAPQ

0.059

1.072

0.565

0.565

1.131

0.026

1.325

0.691

0.723

2.073

0.5

5-H14NAPQ

0.371

1.597

0.807

0.453

2.421

0.209

7.043

1.640

0.685

13.122

0.2

1,2-ACNAQ

0.027

1.747

0.619

0.084

1.858

0.031

0.510

0.208

0.082

0.623

2.9

0.011

0.050

0.030

0.030

0.182

0.028

0.052

0.036

0.030

0.177

1.0

0.019

0.385

0.153

0.125

1.831

0.010

1.031

0.225

0.173

3.145

0.6

0.007

0.068

0.021

0.010

0.107

0.008

0.032

0.020

0.016

0.099

1.1

0.012

0.097

0.060

0.069

0.359

0.011

0.132

0.051

0.030

0.202

1.8

0.015

0.360

0.111

0.033

0.442

0.020

0.102

0.060

0.060

0.241

1.8

0.049

0.411

0.191

0.136

2.865

0.045

0.362

0.181

0.128

2.720

1.0

-

-

-

-

ND

-

1,4-ANTQ 9,10-ANTQ 9,10-PHEQ 2-MANTQ 1,2-ACANQ

EP

AC C

Min

TE D

Max

gas/particle partition

Quinone

7,12-BANTQ

b

1,4-CHRYQ

-

-

-

-

ND

5,6-CHRYQ

0.021

0.119

0.044

0.038

0.349

0.020

0.087

0.048

0.049

0.192

1.8

5,12-NAPACQ

0.320

0.320

0.320

0.320

0.320

0.022

0.082

0.052

0.052

0.104

3.1

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0.078

0.131

0.097

0.081

0.289

0.021

0.088

0.055

0.055

0.109

2.6

4,5-BaPyQ

-

-

-

-

ND

-

-

-

-

ND

-

6,12-BaPyQ

-

-

-

-

ND

-

-

-

-

ND

-

-

-

-

-

(38.007)

-

-

-

-

(74.952)

-

-

-

-

-

12.154

-

-

-

-

22.764

-

(Total w. Excluded) Total w/o Excluded

ND: no analytical signal detected.

c

NA: could not be calculated.

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b

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309

sampling breakthrough in the gas phase.

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Values enclosed in brackets were likely underestimated by some unknown factor because of

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a

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310

Table 3: Nighttime ambient quinones measured in 12-h time-integrated samples (n=15

311

days) Gas-Phase (ng/m3)

Particle-Phase (ng/m3)

Min

Max

Mean

Median

Total

Min

Max

Mean

Median

Total

gas/particle partition

1,4-BQ

(0.367)a

(6.818)

(2.846)

(2.685)

(14.229)

1.438

9.626

3.429

2.151

27.432

NAd

1,2-NAPQ

(5.191)

(5.191)

(5.191)

(5.191)

(5.191)

1.685

1.685

1.685

1.685

1.685

NA

1.941

5.704

NA

0.053

0.865

NA

0.791

0.791

2.1

1.348

5.304

0.3

-

-

-

ND

0.067

3.696

1.901

2-M14NAPQ

(0.010)

(1.195)

(0.207)

(0.029)

(2.689)

0.025

0.122

0.062

2-H14NAPQ

0.533

1.127

0.830

0.830

1.660

0.791

0.791

0.791

5-H14NAPQ

0.320

1.369

0.845

0.845

1.689

0.265

1.699

1.061

1,2-ACNAQ

b

-

-

-

b

0.153

0.153

0.153

0.153

0.153

NA

1,4-ANTQ

0.034

0.140

0.071

0.052

0.424

0.041

0.262

0.106

0.076

0.633

0.7

9,10-ANTQ

0.015

0.257

0.101

0.086

1.420

0.054

0.470

0.174

0.165

2.442

0.6

9,10-PHEQ

0.056

0.333

0.195

0.195

2-MANTQ

-

-

-

-

1,2-ACANQ

0.026

0.044

0.035

0.035

7,12-BANTQ

0.057

0.333

0.162

0.106

1,4-CHRYQ

-

-

-

-

5,6-CHRYQ

0.027

0.085

0.068

0.081

5,12-NAPACQ

b

-

-

-

1,6-BaPyQ

-

-

-

4,5-BaPyQ

-

-

6,12-BaPyQ

-

-

(Total w. Excluded)

-

-

Total w/o Excluded a

-

-

315 316 317

M AN U 0.011

0.078

0.040

0.027

0.201

1.9

ND

0.016

0.032

0.024

0.024

0.048

NA

0.069

0.038

0.055

0.044

0.040

0.132

0.5

2.434

0.063

0.313

0.158

0.107

2.363

1.0

ND

0.041

0.067

0.054

0.054

0.108

NA

0.273

0.021

0.025

0.023

0.023

0.050

5.5

b

-

-

-

-

ND

-

-

ND

-

-

-

-

ND

-

-

-

ND

-

-

-

-

ND

-

-

-

ND

-

-

-

-

ND

-

-

-

(30.467)

-

-

-

-

(47.730)

-

-

-

8.358

-

-

-

-

12.044

-

EP

breakthrough in the gas-phase. b

S/N<10; ND: could not be quantified.

c

ND: No analytical signal detected.

d

NA: could not be calculated.

AC C

314

0.389

Values enclosed brackets are likely underestimated by some unknown factor due to sampling

312 313

SC

-

TE D

1,4-NAPQ

c

RI PT

Quinone

318

When the contents were compared for each phase, the quinone levels in the PM (Figure 2A)

319

were 1.4-1.9 times as abundant as those in the GP (Figure 2B), for the samples that were

320

collected at nighttime and during the daytime. The data analysis confirms that quinone

321

concentrations in the PM exceed those in the GP, whether originating from 12-h (Tables 2 and 3) Page 17 of 40

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or from 24-h samplings (Table 1). Since PM-associated species can travel farther unchanged

323

(Lohmann and Lammel, 2004; Lammel et al., 2009; Keyte et al., 2013), a high proportion in the

324

PM implies that the bulk of quinones originating from precursors emitted during oil sands

325

exploration can be transported over distances to impact distant ecosystems (Harrison et al.,

326

2016). Consequently, current monitoring programs need to be enhanced with new monitoring

327

stations located farther but along the wind trajectory from oil sands emission sources in order to

328

gain a better understanding of quinone dispersion.

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329

Time-resolved concentration profiles demonstrated that quinone levels were most abundant for

331

PM sampled during the day (as shown in Figure 2A) and had the lowest abundance for GP

332

sampled at night. These observations further support daytime conditions as a major venue for

333

oxidized-PAH generation. The daytime and nocturnal variations in the concentration and

334

percentage of the PM quinone content suggest that temperature-controlled partitioning influences

335

the extent of radical-induced secondary reactions and photolysis.

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336

The differing concentration levels amongst quinone positional isomers (Tables 1-3) indicate

338

differences in their generation from the same parent PAH and in their stability. This observation

339

is important because isomers are usually characterized by differing levels of toxicity (CCME,

340

2008; IARC, 2013; Knecht et al., 2013).

342 343

3.2. Effects of the duration and sampling time of day on quinone content in PM

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337

Figure S5 allows for the comparison of the quinone content determined from 24-h PM sampling

345

with the sum of the contents from 12-h daytime and 12-h night PM sampling. In general, the

346

mean of the 24-h PM concentrations was about twice lower than the combined average of day

347

and night PM concentrations. This may imply that a 24-h sampling period is too lengthy and

348

reflects the progress of quinone transformation when compared to 12-h sampling periods (as

349

discussed in the next section).

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The daytime/nocturnal data indicated different patterns in the occurrence of quinones. The

352

highest daytime–nocturnal concentration discrepancy in the total sample (GP+PM) was for 2Page 18 of 40

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MANTQ and 1,2-ACNAQ, which had a daytime concentration that was over 16 times higher

354

than the nighttime concentration (Table S8). This emphasizes the importance of photolysis of

355

precursors and OH radical-initiated reactions to quinone formation (Keyte et al., 2013 and

356

references therein). On the other hand, the 1,4-ANTQ and 9,10-PHEQ nighttime concentrations

357

were three times higher than the corresponding daytime concentrations. This may suggest that

358

NO3 radicals play a significant role in the oxidation processes that led to these two compounds.

359

NO3 has been suggested to be a leading nighttime oxidant by numerous authors (Keyte et al.,

360

2013 and references therein).

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The 12-h daytime/nocturnal collection also demonstrated that, out of 19 quinones included in this

363

study, 5,12-NAPACQ and 1,6-BaPyQ were only detected during the day (Table S8). These

364

quinones had higher contents in the GP than in the PM (Table 2). The results from the 24-h

365

sampling (Table 1) further indicated that 5,12-NAPACQ was mostly present in the PM, and 1,6-

366

BaPyQ was significantly higher in the PM than in the GP. Above observations suggest that these

367

two rather stable quinones were generated through precursor transformations in the GP

368

(photolysis and radicle-initiated reactions) at a much higher rate than parallel secondary

369

processes in the PM (radicle-initiated reactions) and could have been consequently adsorbed onto

370

the PM. Therefore, the 24-h PM content would reflect the sum of the initial concentrations

371

generated in the GP and PM.

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3.2.1. Effects of sampling duration on possible sampling artifacts

374

Secondary organic aerosols are highly complex and dynamic mixtures where primary emitted

375

precursors (like PAHs) are constantly being transformed into progressively more oxidized

376

compounds like quinones and phthalic acids (Keyte et al., 2013). Levels of ambient air quinones

377

reflect a balance between these successive transformations. With longer sampling, the

378

transformations of compounds already trapped in the sampling media will possibly be more

379

extensive, depending on sampling duration, species reactivity towards ozone, concentration of

380

oxidizing agents, particle properties, and meteorological conditions (Liu et al., 2006; Brown and

381

Brown, 2013). Consequently, these processes may reflect in higher summed 12-h day and 12-h

382

night sampling concentrations than those measured in the 24-h samples (like for most of

383

sampling days shown in Figure S5) as oxidation of quinones already collected continues to other

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compounds not measured in this study. Quinone levels from 24-h samplings can also surpass

385

summed day and night 12-h samplings when oxidation of PAHs, quinone precursors, prevails

386

over quinone oxidation (Kamens et al., 1989). Sampling on August 28 could correspond to

387

oxidation of PAHs already collected during 24-h sampling as NO2 levels were exceptionally

388

higher in the afternoon for that day, and which would not be captured by the daytime 6am-to-

389

6pm sampling (as supported by data in Figure S5, and oxidant data in Table S5 for August 28).

390

Comparison of the summed 12-h with 24-h sampling results confirms that longer sampling

391

periods correspond to quinone levels similar to those that would be observed in more aged air

392

mass. However, due to the complexity of these competing transformation processes, it cannot be

393

concluded with certainty in the scope of this study to what extent the 24-h sampling results

394

reflect ambient air quinone oxidation and contribution from “on filter” artifacts. It could have

395

been beneficial to also perform shorter duration samplings to confirm if quinone levels were

396

equivalent but the sampling equipment used in the current study did not allow for these

397

additional tests. The 24-h sampling time is commonly used to accumulate sufficient particulate

398

mass for ambient air composition (NAPS, 1969; US EPA, 1999; European Committee for

399

Standardization, 2008). These oxidation processes of PAHs and oxy-PAHs have been studied

400

previously (Schauer et al., 2003; Liu et al., 2006; Brown and Brown, 2013) and the authors

401

concluded that changing sampling conditions has varying effects on different compounds and

402

that the effect of sampling cannot be entirely isolated from transformations of precursors and

403

products occurring in ambient air. A more detailed investigation of those transformation

404

processes involved in degradation of quinones and their precursors upon sampling and in the

405

atmosphere would be required.

407 408

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3.3. Determination of quinone sources

409

Atmospheric parameters associated with sampling are summarized in Table S5 and in Figure S3.

410

The quinone concentration data were correlated against the monitored meteorological data and

411

the primary pollutant datasets to assess if any relationship exists. The Pearson correlation results

412

(Table S6) indicate that the quinone concentrations determined in the 24-h samples are correlated

413

strongly with NO2 and temperature, and moderately with the NO, O3, and RH data.

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An inverse correlation of quinones with NO and NO2 and positive correlation with O3 suggest

416

differing role of these oxidizing agents in the formation and stability of the quinones. While

417

strong positive correlation of quinones with primary pollutants (SO2, PM2.5, TSP) would suggest

418

common primary emission sources, an observed negative correlation suggests the quinones’ sink

419

role for NOx (i.e., concentration of NOx decreases while quinone’s increases as demonstrated in

420

Figures 1A and 1C). Therefore, the lack of quinone correlation with the primary pollutants and a

421

strong to moderate correlation with oxidizing agents suggests that the detected quinones did not

422

originate from primary emission sources, but rather were generated through secondary chemical

423

processes and photooxidation of primary emitted PAHs. Unfortunately, solar irradiation data

424

was unavailable to further assess any correlation, but a strong link between solar irradiation and

425

PAHs oxidation products was confirmed in other studies (Lampi et al., 2001; Lampi et al., 2006).

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426

Derived correlation data seem to support gas-phase free radical initiated oxidative

428

transformations of PAHs (Keyte et al., 2013, and references therein). These oxidative

429

transformation processes can alter aerosol chemistry through reactions of nitrogen oxides, OH

430

radicals, and ozone with parent PAHs. Data in Figures 1A, 1B and 1C illustrate that in general

431

higher ambient air temperatures reflect in increased quinone levels (Figure 1A) and decreased

432

PAH and NO2 levels for those days. The observed correlations with oxidizing radicals are in

433

agreement with the observations of others (Vione et al., 2006; Wu et al., 2014; De Laurentiis et

434

al., 2015; Elorduy et al., 2016).

EP

435

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During the sampling period one visible plume episode was recorded on August 23rd. This plume

437

event reflected in concentration spikes in almost all the primary (Figure S3) and secondary

438

pollutants, while the 1,2-NAPQ, 1,4-NAPQ were observed at their highest daily concentrations.

439

Wind direction data was used to further support that the observed plume episode and spikes in

440

primary and secondary pollutants were associated with oil sands activities in the region.

441

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442

Effect of wind direction data (CASA, 2015) on the four different levels of quinone concentration

443

derived for monitoring site AMS13 (Figure S1) indicates that quinone levels were enhanced for

444

air masses originating from south (S), south-west (SW) and west-south-west (WSW) directions

445

from oil sands industry (Figure S6). In addition, quinone highest concentration range (6.51Page 21 of 40

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39.80 ng/m3) correlated with highest mean concentrations of primary pollutants for air masses

447

originating from S, SW and WSW directions, as demonstrated in Figures S7 (quinone correlation

448

with NO2), S8 (quinone correlation with PM2.5), S9 (quinone correlation with TSP), and S10

449

(quinone correlation with SO2). Therefore, the wind directions and correlation amongst

450

monitored pollutants suggest oil sands exploration and processing sites as main GP and PM

451

emission sources. Figures S11 and S7 demonstrate that in addition to southern, also northern

452

winds correlated with significant PM2.5 and NO2 levels which suggests additional sources of PM

453

and GP emissions.

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446

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454

In this study, we have measured constituents of ambient air that was transforming and aging.

456

Since the studied quinones demonstrated higher correlation with oxidizing agents than with

457

primary pollutants, their presence in vicinity of oil sands exploration suggests rapid secondary

458

atmospheric transformations of parent PAHs emitted directly from primary sources. In fact,

459

recent studies have found elevated parent PAH levels in the Athabasca oil sands region

460

proximity (Schuster et al., 2015; Zhang et al., 2015; Hsu et al., 2015, Wnorowski, 2017) that

461

could serve as precursors of more oxidized species like quinones (Bunce et al., 1997; Kwamena

462

et al,, 2006, Atkinson and Arey, 2007).

464 465

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3.4. Quinone gas–particle partitioning

An assessment of the gas–particle partitioning is important for understanding the environmental

467

fate and effects of species. Volatile two- and three-ring PAHs are mostly observed in the GP.

468

Four-ring PAHs are in both phases depending on the temperature and organic matter properties.

469

Heavy five-ring or higher PAHs are usually associated with the PM. Species associated with the

470

PM are thought to be “protected” to some degree by organic matter from further transformations

471

and can be transported farther unchanged, while those in GP are susceptible to accelerated

472

transformations (Lohmann and Lammel, 2004; Lammel et al., 2009). Numerous PAH studies

473

have highlighted that the dependence of the gas–particle partitioning on volatility is a function of

474

the subcooled liquid vapor pressure of species in relation to the temperature (Pankow, 1994;

475

Lohmann et al., 2000; Su et al., 2006). However, as reported by Galarneau (2008) and observed

476

in the present study, other factors may affect this simple partitioning–volatility relationship by

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477

means of physicochemical interactions to varying degrees. For instance, meteorological

478

conditions, physical and chemical changes to organic matter agglomerates, interactions with

479

other species, and the contribution of the non-exchangeable fraction may result in deviations to

480

the gas–particle partitioning predicted by vapor pressure alone.

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481 482

There are many available derived vapor pressure datasets that can complement the characteristics

483

of the determined species. Galarneau et al. (2006) confirmed that the patterns of species

484

behavior (e.g., phase partitioning) are valid regardless of the source of the vapor pressure dataset.

SC

485

A comparison of the vapor pressures of quinones derived from EPISuite™ (Chemspider, 2015)

487

with their analogous PAHs (Table S1) indicated higher vapor pressure values for the PAHs. This

488

implies that the analogous quinones are less volatile than their parent PAHs. A positive

489

correlation (Figure S12A) was also found between the vapor pressures (Table S1) and GP/PM

490

partitioning ratios (Table 1). The derived correlation from the 24-h sampling demonstrates the

491

following: (1) species with vapor pressures below 8 × 10-3 Pa and a GP/PM ratio below 1 were

492

primarily partitioned to the PM; and (2) species with vapor pressures above 1 × 10-2 Pa and a

493

GP/PM ratio above 1 were mainly partitioned to the GP (9,10-PHEQ) (Figure S12A). A distinct

494

relationship between the partitioning and vapor pressure or molecular weight has been reported

495

by others (Simcik et al., 1998, Delgado-Saborit et al., 2013; Alam et al., 2014). These authors

496

observed that the gas–particle phase distributions of PAHs and quinones directly emitted from

497

primary sources follow a sigmoidal response curve that can be predicted from their vapor

498

pressures and molecular weights. However, the results in the present study demonstrate that

499

molecular weights of quinones that are generated through secondary atmospheric transformations

500

are not unequivocally indicative of the phase partitioning (Figure S12B), which precludes such a

501

prediction.

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AC C

502

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486

503

Phase partitioning data from the 12-h sampling indicated a higher proportion of quinones in the

504

GP (Tables 2 and 3) than that derived from 24-h sampling (summed content 3.9 ng/m3 in Table

505

1), particularly in the daytime measurements (summed content 12,1 ng/m3 in Table 2). This

506

observation implies that the bulk of quinones could have been initially generated in the GP at a

507

higher rate during the daytime than at night. As observed with 24-h longer sampling time, their Page 23 of 40

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content significantly decreased in the GP and increased in the PM, likely reflecting quinone

509

adsorption onto particulate matter and concurrent quinone oxidation in the gas-phase. Since

510

partitioning of GP-generated quinones differs initially from that of quinones emitted directly

511

from primary sources (Delgado-Saborit et al., 2013), high levels of heavier quinones in the GP

512

(Tables 2 and 3) may indicate an offset of PAH oxidative transformations in secondary organic

513

aerosols. For 12-h sampling measurements in the present study, secondary oxidation products

514

likely did not reach the partitioning equilibrium, and complex factors controlled their levels in

515

each phase. However, further debate would require more data and is outside the scope of this

516

study.

SC

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518

3.5. Quinone size distribution profiles and oil sands fingerprint

519

Table S9 lists the normalized quinone size distribution data as a function of the partition phase,

521

sampling time, and duration. The PM phase data indicated that lower-mass quinones were

522

predominant in the 12- and 24-h samplings. As the number of quinone rings increased, their

523

proportion in the PM decreased. In addition, three-, four-, and five-ring quinones associated with

524

the PM were characterized by similar contents whether collected over 12- or 24-h. However,

525

PM sampling for 24-h indicated a notably high proportion of one-ring quinone at 71%

526

accompanied by a lower content of two-ring quinones, in contrast to PM sampling for 12-h.

527

Although no quantitative data could be established for corresponding GP sampling results, a

528

similar trend was observed where a higher content of GP 1,4-benzoquinone was accompanied by

529

a lower content of all other quinones over time. Therefore, the comparison of GP and PM

530

contents collected during 12- and 24-h sampling suggests that multi-ring quinones in GP

531

transform after at least 12-h. It was also observed that oxidized three-, four-, and five-ring PAHs

532

appeared more stable when associated with the PM. The observed stability increasing with the

533

molecular weight has also been reported by others (Keyte et al., 2013 and references therein).

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The stability of the heavier quinones in the 24-h PM samples justifies the use of their congener

536

ratios for fingerprint characterization. The diagnostic ratio derived from PM contents originating

537

from oil sands was 0.2 for 9,10-PHEQ/(9,10ANTQ + 1,4-ANTQ), 2.49 for 7,12-BANTQ/(1,4-

538

CHRYQ + 5,6-CHRYQ), and 4.98 for 7,12-BANTQ/1,6-BaPyQ. Delgado-Saborit et al. (2013)

539

calculated ratio was 4.69 for PHEQ/(ANTQ) and 6.25 for 7,12-BANTQ/1,6-BaPyQ at a London

540

traffic site. Thus, the oil sands PM demonstrated distinctly lower 9,10-PHEQ levels than

541

ANTQs compared to an urban traffic site.

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The quinone profiles in the current study can be used as a signature of potential sources and a

544

reference point for future quinone monitoring studies and evaluations of emissions originating

545

from oil sands exploration, bitumen, and possibly petroleum processing. Although the shape of

546

the fingerprint can be affected by factors that control quinone formation and stability, such as the

547

precursor emissions, atmospheric oxidants, organic matter properties, and meteorological factors,

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a comparison of the fingerprints can still be informative. Page 25 of 40

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4. Conclusions

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This study provides the first insights on quinone characterization in ambient air affected by oil

553

sands industry emissions. It contributes to knowledge of secondary aerosol formation and

554

propagation, as well as the effects of duration and sampling time on quinone occurrence. The

555

evidence presented in the current study confirms that oil sands processing leads to the formation

556

of quinones in secondary organic aerosols. Because they are less volatile than their precursors,

557

the bulk of quinones are associated with PM and may travel farther unchanged to affect distant

558

ecosystems. Given their presence in the PM, toxicity, and persistence, quinones should become

559

a key component to monitor in air like PAHs, especially at industry influenced sites. This will

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contribute additional composition information on monitored aerosols for risk assessments. The

561

presence of oxidatively modified PAHs can also serve as an indicator of the aerosol oxidation

562

state and aging.

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ACKNOWLEDGMENTS

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The authors would like to extend their gratitude to the NAPS, labs at River Rd, Organics

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Analysis Laboratory team, and WBEA for sampling, sample preparation and providing

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supplementary data, as well as CARA for funding. We also thank ECCC scientists J. Liggio, T.

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Harner, N. Jariyasopit, C. Austin, E. Dabek and R. Turtle for their critical review and comments

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on this manuscript.

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